Microbial Communities in Coastal Sediments This page intentionally left blank Microbial Communities in Coastal Sediments Structure and Functions

SALOM GNANA THANGA VINCENT Professor, Department of Environmental Sciences, University of Kerala, India TIM JENNERJAHN Senior Scientist & Group Leader, Working Group Ecological Biogeochemistry, Leibniz Centre for Tropical Marine Research, Bremen, Germany KUMARASAMY RAMASAMY Director, Faculty and Academics, SRM Institute of Science and Technology, India Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2021 Elsevier Inc. All rights reserved.

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Introduction vii

1. Source and composition of organic matter and its role in designing sediment microbial communities 1 1.1 Introduction 1 1.2 Organic matter in coastal sediments 2 1.3 Source of organic matter: autochthonous and allochthonous 5 1.4 Quality of organic matter in sediments 10 1.5 Microbial degradation of organic matter 12 1.6 Role of organic matter in designing sediment microbial communities 17 1.7 Microbial diversity and in coastal sediments 19 1.8 Diversity of archaeal communities 30 References 34

2. Sources, types, and effects of nutrients (N and P) in coastal sediments 47 2.1 Introduction 47 2.2 Nutrient sources of coastal 50 2.3 Nutrient enrichment: forms and types 59 2.4 Effect of hypernutrification 61 References 73

3. Environmental variables and factors regulating microbial structure and functions 79 3.1 Introduction 79 3.2 Spatial and temporal heterogeneity 80 3.3 Geological factors 82 3.4 Hydrological factors 85 3.5 Physicochemical factors 85 3.6 Biological factors 95 3.7 Nutritional factors 105 3.8 Natural and anthropogenic disturbances 106 3.9 Presence of contaminants/toxic substances 107 References 108

v vi Contents

4. Biogeocycling of nutrients (C, N, P, S, and Fe) and implications on greenhouse gas emissions 119 4.1 Introduction 119 4.2 Biogeocycling of nutrients 120 4.3 Greenhouse gas dynamics in coastal ecosystems 133 References 139

5. Biodegradation and biotransformation of persistent organic pollutants by microbes in coastal sediments 147 5.1 Introduction 147 5.2 Why persistent organic pollutants? 148 5.3 Anaerobic degradation and pathways 149 5.4 Anaerobic involved 159 5.5 Limitations for anaerobic degradation: electron acceptors 161 5.6 Future prospects 161 References 162

6. Assessment of microbial structure and functions in coastal sediments 167 6.1 Introduction 167 6.2 Culture-dependent methods: the “great plate count anomaly” 168 6.3 Molecular tools used to examine microbial diversity of coastal sediments 169 6.4 High-throughput sequencing technologies 175 6.5 Functional diversity of coastal sediment microbes 177 6.6 Microbial activity in coastal sediment: study of biogeochemical reaction rates in laboratory microcosms 180 6.7 Conclusion and future prospects 181 References 182

Appendix 1: Conclusions and future perspectives 187 Index 189 Introduction

The complex mixture of nutrient and organic matter inputs from marine and terrestrial ecosystems into the coastal zone fuels rich and diverse microbial communities in the sediments, which have an important role in processes. These include mineralization of organic matter and assimilation of nutrients that serve as sources of . The structure (abun- dance, diversity, and taxonomic composition) and functions (metabolic activities, biodegradation, and nutrient cycling) of microbial communities are controlled by multiple factors that vary in space and time. As marine sediments cover more than two-thirds of the earth’s surface and are also regarded as the largest reservoir of organic carbon on earth, the spatial and temporal variations in microbial activity are primarily attributed to the quality and quantity of organic matter. Moreover, the availability of inor- ganic nutrients and terminal electron acceptors also influences the abun- dance and activity of sediment microbes. Hence, Baas Becking’s hypothesis, “everything is everywhere; but the environment selects” is also relevant with regard to sediment microbial structure. Studies on the of sediments have been increasing largely in the recent past owing to the advent of molecular tools for charac- terization and quantification of microbial community structure. The recent progress in knowledge on coastal microbiology provided motivation for this book. A citation network analysis using the keywords “microbial commu- nities—coastal sediments—organic matter” displays the progress and inter- linking of various research topics regarding the structure and functions of coastal microbial communities. A total of 5854 papers grouped in five clus- ters numbered from 0 to 4 different clusters are based on the relevant sub- themes of the subject area. The themes for clusters were ecology and diversity of sediment microbes, organic matter sedimentation and microbial dynamics, microbial and diversity in sediments, microbial inter- actions in sediments, and influence of sediment like bioturba- tion on sediment microbial communities. Fig. 1 represents the most cited research papers in the clusters. Chapter 1, Source and Composition of Organic Matter and Its Role in Designing Sediment Microbial Communities, delineates how the qual- ity and availability of sedimentary organic carbon drive the composition and activities of microbial communities. In coastal sediments,

vii viii Introduction

Biddle Jf, 2006, P N... Parkes Rj, 2000, Hyd... Parkes Rj, 2005, Nat... Parkes Rj, 1994, Nat... Whitman Wb, 1998, P ... Hedges Ji, 1995, Mar... Martin Jh, 1987, Dee... Ploug H, 1999, Mar E... White Dc, 1979, Oeco... Alldredge Al, 1988, ... Hoppe Hg, 1983, Mar ... Alldredge Al, 1986, ... Smith Dc, 1992, Natu... Simon M, 2002, Aquat... Aller Rc, 1994, Chem... Delong Ef, 1993, Lim... Grossart Hp, 1998, A... Kristensen E, 2000, ... Bligh Eg, 1959, Can ... Azam F, 1998, Scienc... Porter Kg, 1980, Lim... Rich J, 1997, Deep-s...

Simon M, 1989, MarPomeroy E... Lr, 2001, Aq... Fuhrman Ja, 1980, Ap... Hobbie Je, 1977, App...

Fuhrman J, 1992, Env... Fuhrman Ja, 1995, Li... Wommack Ke, 2000, Mi... Middelboe M, 1996, A... Noble Rt, 1998, Aqua... Figure 1 Citation network of highly cited research publications relating to microbial communities in coastal sediments. microorganisms create their own microhabitats within which several types of interactions occur within and between communities. Geochemical zonation is a well-documented factor responsible for microbial distribu- tion. However, exceptional reports on the cooccurrence of microbes that occupy different metabolic niches, for example, methanogenic and sulfate-reducing bacteria, contradict the concept of geochemical zonation. Under anoxic conditions, organic matter degradation and remineralization are carried out by the concerted action of diverse microbial communities through a “microbial ” and a “,” which are key processes in coastal sediment biogeochemistry. In the microbial food chain, initially, large substances are hydrolyzed by extracellular enzymes of hydrolytic bacteria, and the organic products released during this reaction are trans- formed further into simple substrates. The major groups of microbial communities according to their functional role in organic matter degradation thus include biopolymer hydrolyzers, nitrate reducers, metal reducers, sulfate reducers, and methanogens. The net activity of this microbial food chain is measured by quantifying the terminal electron accepting processes such as nitrate reduction, iron reduction, sulfate reduction, and methanogenesis. In the microbial loop, the organic matter released by viral lysis of bacteria, archaea, and microalgae is consumed by het- erotrophic bacteria, which, in turn, are consumed by small consumers like flagellates and ciliates. Benthic microbial communities derive benefit from this microbial loop by the transfer of labile organic matter toward the deeper layer of the sediments. Such high-quality organic matter in the sediment layers also causes priming of sediment microbial communities that stimulate degradation of organic matter. Introduction ix

Chapter 2, Sources, Types, and Effects of Nutrients (N and P) in Coastal Sediments, deals with the sources, types, and effects of nutrients like nitrogen (N) and phosphorus (P) in coastal sediments. Natural changes in the coastal environment and particular anthropogenic activities led and still lead to increased nutrient input and eutrophication in many coastal zones. Nutrient input has been increasing globally as a conse- quence of human population increase and the related increased demand for food production and fuel. Obviously, this has led to increased nutrient production and nutrient transport from land toward the marine environ- ment through coastal ecosystems. Sewage and industrial discharges are classified as point sources, whereas terrestrial runoff and atmospheric depo- sition are diffuse sources of nutrient input. The nutrient enrichment of coastal ecosystems results in eutrophication and further in the depletion of dissolved oxygen, thus leading to hypoxia and anoxia in the water column as well as in the sediments. The structure and functions of microbial communities in coastal sedi- ments are regulated by a number of environmental variables and factors vary- ing in space and time, which are highlighted in Chapter 3, Environmental Variables and Factors Regulating Microbial Structure and Functions. They include geological factors such as sediment granulometry, sediment depth, and related substrate availability. In addition, various hydrological processes and physicochemical properties such as pH, , porewater chemistry, redox potential, and availability of electron acceptors and electron donors are also among the determining factors. Biological factors include various types of positive and negative trophic interactions between microbes and the inter- action of microbes with larger organisms in the sediment, for example, bio- turbation. As microorganisms grow rapidly and have short generation times, changes over temporal scales are important. The microbially mediated biogeochemical cycling of carbon and nutri- ents such as nitrogen, iron, and sulfur in coastal sediments is the focus of Chapter 4, Biogeocycling of Nutrients (C, N, P, S, and Fe) and Implications on Greenhouse Gas Emissions. Sediment microbes utilize various forms of these elements as either electron acceptors or donors dur- ing respiration-coupled organic matter assimilation. During this process, the elements are either oxidized or reduced, thus converting them to their respective oxidized or reduced states, including the greenhouse gases car- bon dioxide, , and nitrous oxides. Increased input of nutrients and organic matter due to anthropogenic activities has altered the natural balance of biogeochemical nutrient cycling. For example, increased x Introduction organic matter input and sedimentation lead to increased heterotrophic microbial activity in the sediments, thus resulting in the release of carbon from the sediments in the form of carbon dioxide and methane. Because of these processes, the coastal environment acts as a net source of green- house gases. Coastal sediments are also reported to be repositories of persistent organic pollutants, the use of which is increasing globally. The extent of this pollution in coastal sediments and the role of bacteria, in particular anaerobic ones, involved in the degradation of these recalcitrant compounds in anoxic sediments are discussed in Chapter 5,Biodegradationand Biotransformation of Persistent Organic Pollutants by Microbes in Coastal Sediments. The persistent organic compounds (POPs) comprise a large group of substances with a complex diversity of biologic effects. For exam- ple, chlorinated organic compounds are highly resistant to degradation and are highly lipophilic. However, many microorganisms have evolved mechanisms to degrade these recalcitrant compounds. Although most of them are resistant to microbial attack under aerobic conditions, several anaerobic bacteria that are capable of degrading POPs have been widely reported. Reductive dehalogenation is considered to be the predominant process in the anaerobic transformation of halogenated compounds. In addi- tion to potentially serving as a carbon source, organohalides function as ter- minal electron acceptors in an anaerobic respiration process, termed dehalorespiration or halorespiration. Coastal sediments, because of their inherently anaerobic conditions and abundant carbon and sources, tend to have ample microbial and diversity, which are potentially able to degrade organic pollutants. Unraveling the black box of microbial diversity and characterization of hitherto uncultured microorganisms in coastal sediments, so as to obtain a knowledge of their composition and ecological adaptations, are important. Chapter 6, Assessment of Microbial Structure and Functions in Coastal Sediments, explores the various methods that have been used to assess the microbial structure and function in coastal sediments. Earlier, culture- dependent methods were used; however, with the advent of modern molecular tools using sophisticated instruments, studies on microbial ecol- ogy and diversity have gained momentum in the recent past. Polymerase chain reaction (PCR)based molecular methods consist of nucleic acid extraction, amplification of ribosomal DNA, and analysis of PCR pro- ducts by fingerprinting techniques like, for example, Denaturing Gradient Gel Electrophoresis and Temperature Gradient Gel Electrophoresis Introduction xi

(TGGE), which provide information about the microbial community structure in terms of , evenness, and composition. The metagenomic approach is defined as the production and analysis of shot- gun genomic data from microbial assemblages, which is now used to ana- lyze microbial community structure in diverse environments like coastal sediments. DNA-based stable isotope probing is a method used to detect active microbes in the environment and also to link the taxonomic iden- tity of microorganisms to specific biogeochemical processes based on the principle “you are what you eat.” Exploring microbial diversity and bio- prospecting of novel biomolecules from coastal sediments will certainly go a long way to meet the challenges faced by humanity regarding climate change, pollution, and the sustainable supply of food and energy. This page intentionally left blank CHAPTER 1 Source and composition of organic matter and its role in designing sediment microbial communities

1.1 Introduction As the largest sink for organic carbon (OC), coastal sediments play a key role in the global carbon cycle which, in turn, influences the global car- bon balance and climate change. Quantity and turnover rates of organic matter (OM) are key components of the biogeochemical cycles; they are of great importance to the content and recycling of available nutrients, and are directly related to greenhouse gas emissions (Ni et al., 2001). It is important to note that the quality and quantity of OM in coastal sedi- ments represents a primary nutritional source for the living community (Fabiano et al., 1995; Inglis, 1989; Jedrzejczak, 2002). Primarily, OM pro- vides substrate for the -based food webs that characterize coastal sediments (Raymond and Bauer, 2001). Within the detritus-based food chain, a major component is represented by the OM-bacterial community system that plays a pivotal role in the overall biogeochemical cycles by means of OM , remineralization of OM, and nutrient cycling. Among the microbial communities, bacteria are the primary utili- zers as well as the net mineralizers of OM ( Jedrzejczak, 2002). The special characteristics of bacteria such as their small size, rapid turnover rates, and prompt response to changes in OM input make them suitable candidates to monitor changes in environmental conditions of the sediments. Hence, the study of sedimentary OM and associated bacterial community struc- ture and functions is of primary interest for a better understanding of the ecology, nutrient cycling, and health of coastal ecosystems. This chapter deals with the source and types of OM and role of OM in designing microbial community structure in the sediments. The microbial diversity in coastal sediments across the globe is also discussed.

Microbial Communities in Coastal Sediments © 2021 Elsevier Inc. DOI: https://doi.org/10.1016/B978-0-12-815165-5.00001-7 All rights reserved. 1 2 Microbial Communities in Coastal Sediments

1.2 Organic matter in coastal sediments 1.2.1 Types of sedimentary organic matter Total carbon is the sum of three carbon forms: organic, elemental, and inorganic (carbonates and bicarbonates). Sedimentary OM is of various types (Fig. 1.1).

1.2.1.1 Total organic matter/total organic carbon OM and OC are often confused and used interchangeably. Total organic matter (TOM) includes all the elements such as carbon, hydrogen, oxy- gen, and nitrogen that are components of organic compounds. Hence, total organic carbon (TOC) is different from TOM. OM is not measured

Figure 1.1 Types of organic matter in coastal sediments. Source and composition of organic matter and its role 3 directly and usually measured from OC. OC is converted to OM using a conversion factor of 1.72 (Howard, 1965). Organic matterðÞ% 5 Total organic carbonðÞ% 3 1:72 The OM in sediment pore water is differentiated as particulate organic matter (POM) and dissolved organic matter (DOM) based on size. DOM is generally defined as OM that is smaller and pass through 0.45-μm pore size filters, while, POM is a fraction that is retained in 0.45-μm pore size filters. Transformation of POM to DOM is an essential factor that deter- mines the rate of microbial processes in the sediments.

1.2.1.2 Particulate organic matter/particulate organic carbon POM includes both living organisms such as bacteria, phytoplankton, pro- tozoa, and metazoa as well as particulate detritus. Particulate organic car- bon (POC) varies spatially and temporally due to combined influences of primary , water exchange, sediment resuspension, terrestrial (Fan et al., 2018), and fluvial inputs as well as chemical and microbial transformations during their descent in the water column (Sempéré et al., 2000; Shaffer, 1996). Various processes that affect the fate of POC include zooplankton grazing, physicochemical disaggregation and hydrolytic activ- ities are also important pathways that lead to production of dissolved organic carbon (DOC) (Cho and Azam, 1988; Herndl, 1988).

1.2.1.3 Dissolved organic matter/dissolved organic carbon DOC includes a series of OM from simple organic acids to complex mac- romolecular substances such as humic acid and fulvic acid. DOC origi- nates from recent humus of plant litter and soil OM and considered as the carbon source directly used by microbes in the sediment (Lu and Xu, 2014). Quantitatively DOC represents the most important carbon pool (Emerson and Hedges, 2008) and microbial degradation is an important process that significantly affects the fate of DOC. This eventually leads to production of bacterial biomass, thus influencing the flow of C and energy through microbial as well as the release of CO2 to the atmosphere (Azam et al., 1983; Del Giorgio and Cole, 1998).

1.2.1.4 Dissolved inorganic carbon Important inputs of dissolved inorganic carbon (DIC) are solution of atmo- spheric CO2, inorganic carbon bound in the sediments as well as biological processes such as respiration. Fermentation, though not as energetically 4 Microbial Communities in Coastal Sediments favorable, may be responsible for a large fraction of the DIC resulting from the decomposition of organic-rich sediments (Abell et al., 2009). DIC con- 2 sists of three species that are the bicarbonate ion (HCO3 ), the carbonate 2 ion (CO3 ), and the aqueous carbon or carbonic acid (H2CO3).

1.2.1.5 Labile organic matter The labile or active fraction of OM consists of low-molecular-weight (LMW) compounds such as amino acids, carbohydrates, fatty acids, vita- mins, pigments, and nucleotides that can be readily utilized by microor- ganisms. This fraction has a turnover within 23 years and originates from new residues and living organisms including microorganisms. The ratio of the sum of protein, carbohydrate, and lipid carbon, which is called labile organic matter carbon (LOM-C) to TOC is an index of lability (Gonsalves et al., 2011). The capacity of the sediment to supply nutrients is defined by the proportion of TOM that is labile.

1.2.1.6 Refractory organic matter Refractory organic matter is the passive fraction of TOM that is chemi- cally stable. This slow degrading fraction consists primarily of organic compounds that are resistant to degradation and has a turnover rate of 2040 years.

1.2.1.7 Microbial biomass carbon Microbial biomass carbon (MBC) is the most active section in sediment OM and refers to the internal carbon in live bacteria, fungi, , and soil animals. Although MBC accounts for only a small portion of total carbon, it is considered as the main driving force for decomposition of OM (Lu and Xu, 2014) and important characteristic indicators for expressing the activity OC pool (Liang et al., 1998). MBC is correlated closely with cycling of nutrients such as C, N, P and S.

1.2.1.8 Biopolymeric carbon Biopolymeric carbon (BPC) is defined as the sum of the carbon equiva- lents of total carbohydrates, proteins, and lipids. BPC concentrations were calculated as the sum of protein, carbohydrate, and lipid carbon equiva- lents, using conversion factors obtained from the elemental analysis of 2 standard molecules (0.49, 0.4, and 0.75 μgCμg 1 for bovine serum albu- min, , and tripalmitin, respectively) (Fabiano et al., 1995). Source and composition of organic matter and its role 5

1.3 Source of organic matter: autochthonous and allochthonous Multiple sources of OM, including allochthonous terrigenous materials exported from land by rivers and urban runoff, as well as autochthonous production of algae and intertidal vegetation (Goni et al., 2003) are received by coastal sediments. OM in sediments is derived from several sources: 1. pelagic OC that reaches the sediment before it is degraded, including that produced by ; 2. benthic organism’s detritus; and 3. external OC, for example, runoff from agricultural land and urban runoff. Although the sources of OC have been identified, their relative contri- butions to the total flux into the coastal sediments are not yet determined as a large fraction of riverine OM is mineralized beforehand in the estuar- ies (Abril et al., 2002). Riverine OM from different sources reaches the coastal ocean through estuaries that are highly dynamic and where hetero- trophy dominates autotrophy. As the biologically reactive fraction of the riverine OM is almost entirely mineralized in the estuaries, the contribu- tion of OC from rivers to the coastal oceans may often be overestimated.

1.3.1 Autochthonous organic matter Continental margin sediments play a critical role in the global marine carbon cycle (Berner, 1982). This is due to the fact that most of the oceanic occurs in continental shelf region, although they account only for 8% of the total ocean area (Wollast, 1991). Continental margins account for 83% of the global sedimentary OM mineralization ( Jørgensen, 1983)and for about 87% of the TOC buried in marine sediments (Berner, 1982). The largest source of OC to the coastal sediments is in the epipelagic zone, where photosynthesizing microorganisms convert inorganic carbon into organic molecules in the form of POM. Although coastal waters cover only 7% of the total global ocean, they account for 14%30% of primary production in the ocean that varies spatially and temporally. Since the density of POM is close to that of seawater, it will not sink and hence, most of this POM is remi- neralized in the water column by microbial respiration with resultant produc- tion of CO2. Nevertheless, part of POM is transported to the sediment by gravitational transport as well as active transport by zooplanktons. While most of the net primary production is funneled through DOM into the microbial loop, respiration converts majority of the marine primary 6 Microbial Communities in Coastal Sediments production back to DIC (Del Giorgio et al., 1997). This complex and efficient vertical process of carbon cycling is referred to as “biological pump” that oper- ates ubiquitously in the pelagic zone of oceans (Fig. 1.2). Zooplanktons play a key role in the biological pump as they along with gravity-driven aggregates transport the product of primary production from the euphotic zone to the sediment (Honjo et al., 2008). Dead plant and animal material, POM, and DOM that are transported to marine sediments as aggregates are referred to as marine snow (Alldredge and Silver, 1988; Kiørboe, 2000). This term was named after the visual effect of particles sinking through the water like snow and stores carbon in the sediment for hundreds of years. The size of marine snow parti- cles ranges from .0.5 mm to tens of centimeter in size and in nearshore coastal zones, marine snow consists of detritus, mineral grains, phytoplankton, and microorganisms bound loosely in a mucous matrix. The aggregates of the marine snow are held together by extracellular polymeric substances that are natural polymers excreted by bacteria and phytoplankton in the water column.

Figure 1.2 Autochthonous production of OM (biological pump). Source and composition of organic matter and its role 7

In deep coastal regions, about 80%90% of the sinking OM is reminera- lized by bacteria within the water column. However, in shallow coastal regions, 10%50% of the primary production reaches the sediment (Burdige, 2007). Although both mineralization of terrestrialmaterial(allochthonous)and production of autochthonous material occur simultaneously, in coastal marine environments, the major source is marine together with the detrital nature of OM that implies a negligible contribution of allochthonous and anthropogenic terrestrial input. The autochthonous production of OM depends on the effi- ciency of the biological pump. In high latitudes, where, strong seasonal blooms are dominated by diatoms that enhance primary production resulting in pro- duction of large fraction of OM in the euphotic zone. The OM transported from these systems is highly labile and hence, only a small fraction reaches the sediment as most of the OM is degraded in the water column.

1.3.2 Allochthonous organic matter 1.3.2.1 Transport by rivers The quantity of carbon transported by rivers is an important and well- documented component of the global carbon cycle (Fig. 1.3). In the global

Figure 1.3 Source and types of allochthonous OM in the sediments. 8 Microbial Communities in Coastal Sediments carbon cycle, rivers have a critical role in connecting terrestrial, oceanic, and atmospheric carbon reservoirs. The flux of riverine carbon to ocean is 0.9 Gt 2 Cyear 1 of which about 40% is organic and 60% inorganic (Meybeck, 1993). The organic material transported by rivers to the ocean is important to coastal heterotrophic organisms, although riverine OC represents only a small fraction (0.9%) of net global terrestrial primary production (Zhao and Running, 2010). River water transports atmospheric and terrestrial carbon to the ocean. Most fluvial organic material comes from soils and terrestrial plants (Hedges, 1992). Soils are the major source of OC in coastal sediments, fol- lowed by rocks, with river phytoplankton being negligible at the global scale (Meybeck, 1993). Of the terrestrial portion, DIC and particulate inorganic carbon (PIC) are associated with the weathering of carbonate and silicate minerals as given in the following equations:

1 1 2 21 1 2 Carbonates:CaCO3 CO2 H2O Ca 2HCO3

1 1 2 2 1 21 1 Silicates:CaSiO3 2CO2 3H2O 2HCO3 Ca H4SiO4 Two approaches are available for estimating global fluvial carbon fluxes. One uses carbon data for large rivers in various regions. The other approach considers the mass balance. The main factors that govern POC fluxes are the total mass of suspended and sediment load (Huang et al., 2017). Rivers export around 0.25 Gt of DOC and 0.15 Gt of POC from continents to the ocean every year (Hedges et al., 1997). The main determinants of DOC fluxes are the drainage intensity, basin slope, and amount of soil OC. For example, the POC flux from rivers may increase exponentially following extreme events such as a typhoon, which is a high-intensity, low-frequency event (Goldsmith et al., 2008). The major sources of riverine OC are living and dead biomass that is contributed by varying concentrations of DIC (35%45%), DOC (22%29%), POC (21%26%), and PIC (17%) (Galy et al., 2015; Meybeck and Vörösmarty, 1999). The tropical region (23.5°N23.5°S) includes 42.7% area of the world’s land; however, it contributes to a disproportionate 66.2% of global freshwater outflow, 73.2% of sediment load, and over 61% of ter- restrial net primary production (Syvitski et al., 2005). For example, in tropical Asia, and particularly in Indonesia, the unusually high DOC con- centration (2200 μM) is characteristic of reported black-water rivers and is caused by their basins mostly covered by peat (Huang et al., 2012). The source of high DOC concentrations in the Nyong River, Africa, is mostly Source and composition of organic matter and its role 9 plants and kaolinite, which are rich in old OM, in the river basin (Brunet et al., 2009; Olivie-Lauquet et al., 2000). Huang et al. (2012) reported mean discharge-weighted DOC concentrations of 616, 411, 431, and 399 μM in tropical Africa, the Americas, Asia, and Oceania, respectively. Therefore tropical rivers are critical to total global fluvial carbon flux. Unfortunately, prior studies for global fluvial carbon fluxes are incomplete because only a few large tropical rivers are considered (Huang et al., 2017) in detailed studies.

1.3.2.2 Agricultural and urban runoff In addition to the natural fluvial transport by rivers, OM derived from domestic, agricultural, and industrial wastes that are difficult to quantify can significantly contribute to the global carbon budget in coastal sedi- ments (Meybeck, 1993). Changes in land use and discharges of sewage that are related to an increase in the transport of OC from land have con- tributed to increase in the degree of heterotrophy in the coastal zone and in the accumulation rate of OC in coastal sediments during the last cen- tury (Ver et al., 1999). This might be the case in many countries, espe- cially in developing countries, where sewage treatments are lacking. It is also shown that the introduction of sewage treatments in densely popu- lated basins results in a decrease of the OC concentrations close to the natural levels (Abril et al., 2002). Moreover, the enhancement of soil erosion by agricultural processes increases POC input (Ciais et al., 2008) in coastal ecosystems. Terrestrially derived OM is relatively more refractory than autochthonous OM due to the presence of more resistant biopolymers such as cellulose, lignin, and cutin. The refractory nature of terrestrial OM is also due to degradation and ageing it undergoes during burial and transport. Increased inputs of refractory POC from soils due to changing land-use practices would decrease the mineralization efficiency in estuaries. Nevertheless, discharge of sewage containing highly labile POC would increase the mineralization rate. Hence, soil erosion would result in the increase of net POC transfer to coastal sediment, whereas sewage discharge would limit the net POC transfer (Abril et al., 2002). Moreover, long residence time estuaries behave as filters and buffer the increased carbon load by efficient minerali- zation. Ultimately, the nature of perturbation (sewage discharge) induces increased burial of OC in coastal sediments, enabling them to act as potential carbon sink. Anthropogenic sewage input is a major source of 10 Microbial Communities in Coastal Sediments

DOC in coastal sediments, with increased consumption in the upper reach low salinity zones (He et al., 2010).

1.4 Quality of organic matter in sediments The quality of OC is an important factor influencing the degradation rates and incorporation efficiency. The allochthonous OM entering the coastal system either by fluvial flow or terrestrial runoff may be refractory or labile depending on the source. Refractory OM relates to fresh OM, whereas, OM from sewage and dying freshwater algae and bacteria is predomi- nantly labile. The OM, while being transported downstream from riverine to marine zone undergoes decomposition and is not further supplemented with fresh labile organic matter (LOM) because autochthonous produc- tion is limited in the marine part (Middelburg et al., 1993). The quality of OM expressed by its first-order decay constant decreases upon mineraliza- tion because the labile substrates are consumed initially and consequently, the remaining substrate becomes more refractory (Westrich and Berner, 1984). This ageing effect has been observed in several environments cov- ering various timescales (Middelburg, 1989). Apart from this, local primary production results in a slight increase of LOM and hence, OM mineraliza- tion rates (Middelburg et al., 1993). Riverine DOC contains a highly var- iable, but significant labile fraction that can be consumed at the timescale of estuarine mixing. The existence of simultaneous sources and sinks in estuaries results in apparent conservative behavior of DOC and hence results in negligible net changes in bulk concentrations (Abril et al., 2002).

1.4.1 Organic matter quality indices The state of OM degradation is estimated using several maturity indicators that vary from short-term (e.g., chlorophyll) to long-term (e.g., nonpro- tein amino acid) indicators. Biopolymers such as carbohydrates and proteins contribute to major part of OM contributing to 3%30% and 2%15%, respectively, of the bulk DOC in coastal surface water. These LOM fractions play important roles in OM cycling in the sediments. Moreover, in benthic ecology studies, nitrogen-containing compounds such as proteins (amino acids) and hexosamines (amino sugars) (AA and HA) can be used as potential indicators as they regulate and limit hetero- trophic metabolism (Tenore, 1983) and amino acids reflect the degrada- tion rate of OM. Jennerjahn and Ittekkot (1997) proposed the reactivity index as an indicator of OM reactivity. This is based on the concept that Source and composition of organic matter and its role 11 fresh OM contains more of reactive aromatic amino acids such as tyrosine and phenylalanine that are rapidly degraded. More degraded OM is enriched with amino acids such as glycine, serine, and threonine, whereas, increasing degradation of OM depletes certain amino acids such as phe- nylalanine, glutamic acid, tyrosine, leucine, and isoleucine. This degrada- tion pattern based on molecular composition of protein amino acids was used to derive a quantitative Degradation Index (Dauwe and Middelburg, 1998). Amino sugars and amino acids contribute to the labile component of OM and they are good indicators of freshness of settling POM in the marine sediment. Being easily degradable component, AA and HA are consumed by microbes in the water column during sedimentation. Moreover, combinations of AA and HA are used to understand the bio- geochemical nature of OM in the sediments (Gupta and Kawahata, 2003). Apart from the earlier mentioned indices, ratio of LOM-C to TOM-C is also considered as the lability index, which reflects the OM stability. The OM stability would be lesser with higher proportion of LOM-C/TOM-C (Lu and Xu, 2014). The carbohydrate fraction of LOM-C has complex chemical structures that resist extracellular enzymatic hydrolysis and persist in dissolved forms (Arnosti and Holmer, 1999). This fraction forms part of uncharacterized DOM in the sediments and consists of LMW compounds (Hedges et al., 2000; Keil et al., 1994; Benner et al., 1992). Monomeric sugars are easily remineralized, while more complex structures such as cellu- lose are more refractory. Certain carbohydrates resist microbial degradation that explains their persistence and hence their concentrations either remain relatively constant or change slightly with depth. The (Protein:Carbohydrate) PRT:CHO ratio gives a measure of the quality of OM (Danavaro et al., 1994; Fabiano et al., 1995) and a low PRT:CHO ratio together with the low concentrations of BPC relates to oligotrophic conditions. Among the three fractions, carbohydrate usually constitutes the dominant fraction and the contribution of carbohydrates to the DOC pool is substantial. This indicates a detrital or allochthonous ori- gin of OM (Fabiano et al., 2004). Nevertheless, changes in the quality of OM occur both spatially and temporally. During summer, rapid con- sumption of the labile fraction (proteins and lipids) leads to an accumula- tion of the refractory component (carbohydrates) (Fabiano et al., 2004). However, proteins were reported to be the major component of the LOM in a tropical coastal ecosystem with an average of 71% followed by carbohydrates 20% and lipids 9% (Gonsalves et al., 2011). Total bacterial 12 Microbial Communities in Coastal Sediments density showed a high correlation with carbohydrates, while no significant correlation was found with the lipid fraction. Thus proteins and carbohy- drates were the main organic compounds affecting the bacterial density in coastal sediments. In addition, the significant positive correlation observed between the frequency of dividing cells and the PRT:CHO ratio showed that the biochemical composition of OM also has an influence on the active bacterial fraction (Fabiano et al., 2004).

1.5 Microbial degradation of organic matter OM within coastal and marine sediments originates from the sedimenta- tion of dead plant and animal material from the water column. Hence, the nature of the deposited material and the chemical, biological, or phys- ical processes that affect this material after its deposition determines the geochemistry of marine sediments (Burdige, 2006). Among the various pathways, microbial degradation of OM is an important source of OC in the sediments. Organic compounds and associated nutrients supplied to sediments are mineralized through heterotrophic decomposition (Capone and Kiene, 1988; Megonigal et al., 2004). These processes are referred to as early diagenesis and are mainly mediated by bacteria. Moreover, sub- oxic and anoxic bottom water in coastal regions with high primary pro- ductivity result in the formation of anoxic sediments, since most of the oxygen is consumed within the first few millimeters of the sediment dur- ing OM mineralization (Canfield, 1993). Therefore most of the decom- position of OM in organic-rich coastal sediments is anaerobic, because these sediments often become anoxic close to the sedimentwater inter- face (Burdige, 2006). In microbial ecology, the feeding niche of a is defined by three factors: (1) the electron source or donor, (2) electron sink or ter- minal electron acceptor (TEA), and (3) carbon source. Those bacteria that use light energy are phototrophic and those that use chemical energy are termed chemotrophic. When CO2 is used as a carbon source, the organ- isms are termed autotrophs and when an OC source is used, they are het- erotrophs. When the electron source is inorganic, the organisms are named and those using organic compounds as electron sources are defined as (Table 1.1). Microbes derive energy by transferring electrons from an external electron donor to an external TEA (Fig. 1.4). The electron donors may be organic compounds such as acetate, CH4, or simple monomers (fermentation) and Source and composition of organic matter and its role 13

Table 1.1 Classification of microbes based on energy, carbon, and electron source. Energy Carbon Electron Microbial classification Light Organic Inorganic Photolithoheterotrophs Light Organic Organic Light Inorganic Inorganic Photolithoautotrophs Light Inorganic Organic Photoorganoautotrophs Chemical Organic Inorganic Chemolithoheterotrophs Chemical Organic Organic Chemoheterotrophs Chemical Inorganic Inorganic Chemolithoautotrophs Chemical Inorganic Organic Chemoorganoautotrophs

Figure 1.4 Microbial degradation of organic matter.

1 inorganic electron donors are H2, ammonium (NH4 ), (H2S), etc. Fermentation is a metabolic process in which the organic com- pounds serve as both electron donors and acceptors. During the process of electron transport within a cell, from the donor to acceptor, energy is har- nessed by the bacteria. The electron acceptors are oxygen (O2)foraerobic 2 bacteria; nitrate (NO3 ), manganese (Mn[IV]), and ferric iron (Fe[III]) for facultative anaerobes; sulfate and carbon dioxide for obligate anaerobes such as sulfate-reducing bacteria (SRB) and methanogens, respectively. 14 Microbial Communities in Coastal Sediments

In the sediments, microbes degrade OM using a sequence of respira- tory and fermentative metabolisms. Initially, cellulolytic bacteria hydrolyze organic polymers through extracellular enzyme production and further breakdown monomers to alcohols, fatty acids, and hydrogen (H2) through fermentation (Fig. 1.5). By respiration, bacteria gain more ATP through glycolysis, electron transport chain, and oxidative phosphorylation. The energy yield from fermentation through glycolysis and substrate-level phosphorylation is much lower; however, is sufficient for growth of the organisms ( Jørgensen, 2000). This means that the maximum free energy gain for microorganisms is associated with aerobic respiration and the minimum gain is associated with methanogenesis. Although fermentation is not energetically favorable as that of respiration, it may be responsible for large fractions of DIC efflux in organic-rich coastal sediments (Abell et al., 2009). Complete oxidation of a broad range of organic compounds in these systems can occur through the sequential activity of different

Figure 1.5 Anaerobic degradation of organic matter. Source and composition of organic matter and its role 15 groups of anaerobic bacteria (Capone and Kiene, 1988). Fermenting microbial communities is important in the first steps of OM degradation and studies of anoxic tidal sediments have revealed that bacterial commu- nities in the sediment column are dominated by fermentative bacteria (Köpke et al., 2005; Wilms et al., 2006) as well as being typically found also in other anoxic sediments (Schink, 2002). The products of hydrolysis are subsequently consumed by fermenting and acetogenic bacteria that produce LMW-DOC compounds that include volatile fatty acids (VFAs) such as lactate, acetate, formate, propio- nate, butyrate, valerate, and other end products such as alcohols, aromatic compounds, and hydrogen (Middelburg et al., 1993). Alcohols and fatty acids are then further degraded by syntrophic bacteria (secondary fermen- ters) into acetate, H2, and CO2, which can be used as substrate by metha- nogens (Zinder, 1993; Conrad, 1999). POC in the sediment consists of complex biopolymers that need to be hydrolyzed by extracellular enzymes into dissolved molecules with a molecu- lar mass smaller than 600 Da in order to allow assimilation by bacteria (Weiss et al., 1991). Some of the DOC is rapidly mineralized; however, a part of the DOC is recalcitrant and resists rapid microbial degradation (Williams and Druffel, 1987; Arnosti, 2000). Anaerobic bacteria that ferment carbohydrates assimilate only about 10% of the substrate carbon (Clark, 1989)andexcrete the remainder as LMW organic compounds such as short-chain acids, alco- hols, and CO2. The acids and alcohols are taken up by other bacteria in the community that oxidize much of the carbon to CO2. Oxidationreduction reactions determine the redox couple that is energetically more favorable, and pathways with a high-energy gain are preferentially used over pathways with a lower gain, and in marine sedi- ments the availability of electron acceptors is limited, which results in a zonation of the degradation pathways. Thus, in water-logged sediments with a relatively high OC input, oxygen becomes depleted and anaerobic conditions invariably arise where the anaerobic degradation of organic inputs from the water column produces CO2 and CH4 as the terminal output. The various processes involved in the biodegradation of OM involve complex communities of microbes and there is much to be resolved regarding the succession of microbes in these processes. The metabolic pathway and microbes involved are affected by (1) the quality and degradability of the OC and (2) the availability of TEAs. These are the major factors governing the type of microbes found and they can be grouped into functional categories according to metabolic 16 Microbial Communities in Coastal Sediments pathways: biopolymer hydrolyzing, primary fermenters, secondary fer- menters, nitrate reducers, metal reducers, sulfate reducers, and methano- gens. The primary fermenters produce short-chain fatty acids such as lactate, which is in turn utilized by the anaerobic respirers such as the sul- fate reducers, and acetate is produced, which is further utilized by iron(III) reducers (Lovley, 2006). In the absence of any oxidants, OC mineralization through fermenta- tive processes results in the production of equal amounts of carbon diox- ide and methane according to

2CH2O-CO2 1 CH4 The availability of oxidants results in a proportionally greater amount of carbon dioxide relative to methane either directly or indirectly through oxidation of methane formed by fermentation

CH2O 1 oxidant-CO2 1 reaction products

CH4 1 oxidant-CO2 1 reaction products The reactivity of OM is usually expressed in terms of a first-order rate constant R 5 kC where R is the rate of OM mineralization, C is the concentration of OM, and k is the first-order rate constant. For sediments receiving a certain amount of OC, the residence time or the time to reach 95% of steady-state concentrations can be calculated according to lnðÞ 0:05 3 Residence time 52 5 k k The rate of OM remineralization within the sediment is determined by (1) the amount and reactivity of the available organic substrate, (2) sup- ply of oxidants, (3) the composition and activity of the prevailing micro- bial community, (4) temperature, and (5) sediment fabric (Zabel and Hensen, 2002). Mineralization efficiency increases when increasing the POC input. Relatively low mineralization efficiencies can be attributed to the presence of terrestrial soil refractory POC (Abril et al., 2002). The metabolic quotient (qCO2: proportion of basal respiration per microbial biomass) can be used to indicate ecological efficiency of the soil Source and composition of organic matter and its role 17 microbial community (Degens, 1998). This index is based on Odum’s theory of ecosystem succession (1969), where during ecosystem succession toward maturity, there is a trend of increasing efficiency in energy utiliza- tion concomitant with an increase in diversity. A high qCO2 indicates inefficient use of energy, while a low qCO2 indicates high efficiency and more carbon utilized for biomass production (Francaviglia et al., 2004). Microbial degradation of OM consists of two first-order reactions with different rate constants, which include the decomposition of labile and refrac- tory material (Newell et al., 1981). The OC within the deeper sediments where the rate of sulfide formation exceeds its oxidation rate is more resistant to further mineralization and sustains the low rate of sulfate reduction over great sediment depths ( Jørgensen, 1982). LMW-DOC that resists degradation accumulates with depth in sediment pore water (Burdige and Gardner, 1998; Amon and Benner, 1996; Hedges, 1988). Hence, pore water DOC concen- trations reflect the balance between rates of production and consumption (Alperin et al., 1999). In the deeper sediment where sulfate reduction rates decrease, low VFA concentrations will be maintained by methanogenic bac- teria in the sulfatemethane transition (SMT) zone. Ultimately, it is important to determine the amount of OM, which is mineralized relative to the amount, which is preserved in marine sedi- ments, since this affects the levels of CO2 and O2 over geological time- scales (Berner et al., 2007). For this purpose, a better understanding of the processes that control OM degradation in marine sediments is required. Specific emphasis should be placed on continental margins, with the focus on the flux of OM, and the degradation and accumulation of OC in sedi- ments with high productivity. On the one hand, increased degradation of sedimentary OC decreases the carbon storage in sediments resulting in release of more carbon into atmosphere, thus causing increase of atmo- spheric carbon dioxide concentration (Huo et al., 2013; Lu et al., 2014). On the other hand, global warming could accelerate decomposition of soil OM and release of carbon into atmosphere, which would further strengthen the trend of global warming ( Jenkinson et al., 1991).

1.6 Role of organic matter in designing sediment microbial communities Although the dynamics of OM in coastal sediments have been extensively studied and despite its global importance, questions still remain regarding the sources, fate, and role of sediment OC in the microbial diversity and 18 Microbial Communities in Coastal Sediments functions. The reasons for these uncertainties include the complex interac- tions among the various physical, geological, and biological factors that define each estuarine system and control OM cycling in these environments. The linkage observed between sea and land is the main factor controlling the ori- gin and nature of OM in coastal sediments as well as regulating bacterial community structure (Fabiano et al., 2004). Consequently, sediments with different biogeochemical properties have microbial communities that exhibit distinct catabolic responses to a range of carbon sources. Thus the composi- tion and activities of microbial communities are regulated by the quality and availability of carbon. Microbial response to the OM in the sediment depends on the physio- logical capabilities of the microorganisms (Madrid et al., 2001) that ulti- mately determine which microorganisms can coexist successfully within the environment. In several ecosystems including coastal ecosystems, microbial communities form biological organizations, which are horizon- tally stratified and referred as “microbial mats” (Prieto-Barajas et al., 2018). They comprise microorganisms that are highly diverse and abun- dant and also interact ecologically by exchange of signals and also by forming different type of associations. These microorganisms network with each other by coupling biochemical processes and also drive the bio- geochemical cycles (Fig. 1.6). The dominant microbial populations vary with depth and interact with other populations in the efficient cycling of carbon and nutrients (Canfield et al., 2005). In coastal sediments, variations in OC substrate variability with depth produce vertical gradients, presenting niche variability, which dictates

Figure 1.6 Vertical microbial community structure and microbial mats. Source and composition of organic matter and its role 19 how the available energy resources are distributed, resulting in niche diversification by microorganisms. Julies (2007) proposed that organisms with copiotrophic characteristics thrive in the surface 6 cm of the sedi- ment, whereas oligotrophic-type organisms thrive in the deeper sedi- ments. According to the “niche overlap hypothesis” (Pianka, 1974), highly diverse communities arise in environments, which are stable over long periods of time as a result of -maintained niche diversifi- cation. Microorganisms that are adapted to thrive using readily available carbon sources, which occur in the top 2 cm of the sediment, are called and they are “zymogenous,” which means they have the ability to ferment carbohydrates (Koch, 2001). Bacteroidetes bacteria can be clas- sified as copiotrophs and are dominant in environments with high carbon availability and remineralization rates (Fierer et al., 2007). Active members of Bacteroides were previously detected in anoxic sediments (Mouné et al., 2003; Ravenschlag et al., 2001) and this group contains bacteria with hydrolytic and fermenting abilities (Weller et al., 2000). These microorganisms exhibit r-strategists life histories, and have high growth rate that agrees with the higher abundance of bacteria in the top 6 cm. In contrast, in the deeper sediment (1012 cm depth) where DOC becomes increasingly unavailable to microorganisms, species richness is high and bacterial abundance is low. This is because the higher amount of refractory OC in the deeper sediment makes it an environment with lim- ited carbon resources. Competition for the limited, accessible DOC within the deeper sediment may lead to niche specialization and diversifi- cation, resulting in higher phylogenetic diversity. Microorganisms whose activity are adapted to low amounts of readily degradable carbon sources or to compounds that are not readily biodegradable are referred to as oligo- trophs and they have an allochthonous mode of feeding (Koch, 2001).

1.7 Microbial diversity and ecology in coastal sediments The vast microbial diversity is a product of long evolutionary history, proba- bly 2 billion years more than that of eukaryotic organisms, which explains the reasons for the diverse they occupy. Understanding and quantify- ing microbial remains an exciting and significant challenge in microbiology: “Who’s there, what are they doing and how does this relate to ecosystem processes?” (DeLong, 2009). The importance of microbes also exists in the fact that they are crucial component of the biosphere as they act as catalysts for several biogeochemical processes that sustain life on earth. Baas 20 Microbial Communities in Coastal Sediments

Becking’s long-standing hypothesis “everything is everywhere, but the envi- ronment selects” (Baas Becking, 1934 p. 15, translation according to de Wit and Bouvier, 2006)ishighlyrelevantwhenexaminingthemicrobialcommu- nity composition of these unique . The unit of diversity is species that can be defined as assemblage of strains sharing 70% or more DNA homology (Colwell et al., 1995). From an ecological perspective, species is defined as the organisms occupying the same niche and microbial diversity as the number of species and their relative abundance in a community (Atlas, 1984). Microbial diversity is controlled by a variety of ecological factors and evolutionary mechanisms that work at the population and molecular levels. One major reason for the high genome diversity of microbes in the sediments is their innate capacity to accumulate large numbers of mutations and also due to various molecular mechanisms such as lateral DNA transfer and recombination. This results in a population that represents a mixture of genetically diverg- ing species in the like coastal sediments. It is estimated that .99% of microorganisms observed in nature are not readily cultivated using standard techniques (Amann et al., 1995). However, advancements in the understanding and development of molecular tools such as high-throughput 16S rRNA gene sequencing techniques permit culture- independent analysis of bacteria in soil DNA extracts (Roesch et al., 2007)as well as to assess microbial diversity of mixed natural populations (Stahl et al., 1985). Moreover, there has been an explosion of molecular genetic data that has revealed the presence of many unexpected evolutionary lineages and has also proved that the major contributor to total biological diversity is microbial (Pace, 1997). In coastal sediments, molecular analyses of microbial phyloge- netic diversity have identified several key taxa that comprise a significant por- tion of the community in geographically and environmentally contrasting and dissimilar locations (Teske, 2013; Parkes et al., 2014; Carr et al., 2015). However, many of these taxonomic groups lack cultured representatives, leaving their physiological characteristics largely unknown. In-depth analysis of prokaryotic communities in coastal sediments is crucial in understanding ecosystem functioning.

1.7.1 Diversity of bacterial communities Degradation of organic compounds is carried out by a variety of aerobic and anaerobic bacteria in the sediment and contributes to transformations of carbon, nitrogen, iron, and sulfur, thus converting the coastal sediments Source and composition of organic matter and its role 21 as hotspots for biogeochemical cycling. However, in water-logged sedi- ments enriched with organic compounds, oxygen becomes depleted and anaerobic conditions invariably arise where the anaerobic bacteria pre- dominate. The major type of anaerobic microbes found in coastal sedi- ments according to their functional role in OM degradation is biopolymer hydrolyzers, nitrate reducers, metal reducers, sulfate reducers, and metha- nogens. In addition to the presence of abundant uncultured bacteria in the sediment, the prevalence of complex interactions among microbial communities limits the understanding of the underlying metabolic path- ways in these processes. Both autochthonous and allochthonous OM in the overlying water are processed by the sedimentary microbial communi- ties by a variety of metabolic pathways involving various enzymes. Bacteria involved in OC degradation in the marine environment belong to different microbial communities with different rates of con- sumption of various DOC components (Cottrell and Kirchman, 2000) and different growth rates (Cottrell and David, 2003). Only few studies have explored the relationship between species richness and ecosystem functioning in marine environments (Lin et al., 2006; Mills et al., 2005; Emmerson and Huxham, 2002). Ecosystem functioning in this context refers to the metabolic activities of microorganisms, the transformation of OM, and the flow of nutrients, water, and atmospheric gases. Hence, exploring microbial community structure in relation to metabolic activity improves the comprehension of carbon flow between the initial and ter- minal members of the sedimentary microbial food chain.

1.7.1.1 Hydrolytic bacteria The members of this group are involved in the initial degradation of com- plex polymers such as polysaccharides and proteins into simple monomers such as monosaccharides and amino acids, respectively. These organisms are very crucial in OM degradation because the rate of depolymerization of complex polymers exerts a major control on the overall rate of OM decomposition (Megonigal et al., 2004). Members of the Bacteroidetes and Gammaproteobacteria possess enzymes for polysaccharide hydrolysis and fermentation in sediments (Alderkamp et al., 2007). The higher meta- bolic rates in the surface 68 cm of the sediment are matched by a higher abundance of Bacteroidetes and Gammaproteobacteria compared to the deeper sediment. This is because of the fact that most of the carbon input into the sediment is due to vertical flux of settling OM, providing fresh OM and stimulating microbial activity in the surface sediment. The 22 Microbial Communities in Coastal Sediments majority of the Bacteroidetes phylum is gram-negative bacteria and includes aerobic, microaerophilic (Manz et al., 1996), and anaerobic fer- mentative species (Ravenschlag et al., 2001) that have the ability to degrade complex organic macromolecules (Weller et al., 2000). The distribution of these bacteria throughout the depth of the sediment can give an initial indication of a possible link between hydrolysis rates and the abundance of potential hydrolytic and fermenting bacteria. The propor- tionate decrease in cellular abundance of both the potential hydrolytic bacte- ria suggests that they become increasingly carbon-limited at greater sediment depth, which is consistent with the gradual increase in the C:N ratio of the OM. Apart from Gammaproteobacteria and Bacteroidetes, other common phyla reported to involve in hydrolysis in coastal sediments include the Proteobacteria, Chloroflexi, and Planctomycetes (Parkes et al., 2014). One of the most dominant phyla within subsurface sites is the Atribacteria (Dodsworth et al., 2013), a newly classified phylum. Atribacteria have been proposed as heterotrophic anaerobes, with some lineages predicted to special- ize in carbohydrate or fatty acid fermentation (Nobu et al., 2016). So far, few culture-independent studies have been carried out to access the taxonomic diversity of bacteria in mangrove sediments. Previous studies have shown predominant bacterial phylotypes in mangrove sediments to clus- ter within Proteobacteria, Bacteroidetes, Gemmatimonadetes, Actinobacteria, and Firmicutes (Zhang et al., 2009). Sediments of Tuvem and Divar harbor are major sediment bacterial groups affiliated with the phyla Proteobacteria, Bacteroidetes, Firmicutes, Chloroflexi, Planctomycetes, and Actinobacteria. This study also revealed the existence of bacteria belonging to other phyla such as Acidobacteria, Gemmatimonadetes, and members of the candidate divisions in mangrove sediments. The phylum Gemmatimonadetes has been found earlier in systems with high nutrient input (Li et al., 2006). Molecular investigations in a Chinesemangroveecosystem(Liang et al., 2007)showed that the Gammaproteobacteria-affiliated sequences constituted the largest por- tion of their clone library. When compared to the water columns of ocean (Kirchman et al., 2010), the species richness in mangrove sediments is far greater. A considerable fraction of the low-abundance operational taxonomic units of the so-called “rare biosphere” (Sogin et al., 2006) were responsible for the high diversity observed in mangrove sediments and indicate that they have the potential to become dominant when favorable environmental con- ditions arise. Members of the phylum Actinobacteria are ubiquitous in estuarine and oceanic environments. Nearly 15% of the sequences recorded at Tuvem Source and composition of organic matter and its role 23 and Divar belonged to representatives of the phylum Actinobacteria making it the next most abundant phylum following Proteobacteria. Many Actinobacteria are of economic importance (Ward and Bora, 2006) as they are a source of antibiotics (Kim et al., 2006; Manivasagan et al., 2009). Like Deltaproteobacteria, they play multiple roles in the environment that include degradation of cellulose, hydrocarbons (Harwatietal.,2007), metal oxidation (Bryan and Johnson, 2008; Johnson et al., 2009), and nitrate reduction. The presence and activity of Actinobacteria in these sediments could thus be vital in altering the benthic chemistry. Baker et al. (2015) reported Bacteroides, Nitrospira, and Spirochaetes are capable of hydrolyz- ing complex organic compounds using the hydrolytic pathways in estuarine sediments.

1.7.1.2 Denitrifying bacteria Anaerobic conditions and substrate availability in OC-rich coastal sedi- ments (Krishnan and Bharathi, 2009) favor alternate respiratory pathways such as denitrification. Denitrifiers are aerobic bacteria that utilize either 2 2 NO3 or NO2 as TEA in the presence of little or no available O2. However, denitrification occurs rapidly in the below oxic surface zone in 2 the sediments in the presence of NO3 . Although denitrification occurs in aerobic conditions, denitrifiers switch to anaerobic respiration and the 21 critical O2 concentration for this to happen is approximately 10 μmol L (Seitzinger, 1988). Moreover, denitrification can be limited by the avail- 2 ability of carbon in the absence of O2 as well abundance of NO3 . Sediment carbon influences denitrification by promoting anaerobic condi- tions in situ (Tiedje, 1988). Denitrification is thus a major terminal process 2 in the nitrogen cycle converting NO3 to N2, thus removing the fixed nitrogen from the environment. 1 2- 1 2 1 1 5CH2O 4NO3 2N2 4HCO3 CO2 3H2O Denitrification involves the following steps: Step 1 Step 2 Step 3 Step 4 2 - 2 - - - NO3 NO2 NO N2O N2 During this process, nitrogen is not assimilated into the cell and hence, termed as “dissimilatory nitrate reduction.” However, several microorgan- isms also reduce nitrogen oxides without conserving energy for growth, which is nonrespiratory process and may be either assimilatory or dissimi- latory. Denitrification is thermodynamically favorable and predominates other respiratory metabolic processes such as Fe(III) reduction and sulfate 24 Microbial Communities in Coastal Sediments reduction or methanogenesis when nitrogen oxides are available in the sediments. Respiratory denitrification is a versatile anaerobic metabolic process and widespread among several taxonomic groups of bacteria. Denitrification reactions are catalyzed by membrane-bound enzymes: 2 2 NO3 reductase, NO2 reductase, NO reductase, and N2O reductase in the steps 1, 2, 3, and 4, respectively (Fig. 1.7). Microbial nitrate reduction 2 is a dissimilatory process, in which NO3 is reduced with various electron donors by energy-gaining metabolism in the absence or near absence of O2. The gene encoding for N2O reductase (nosZ) is largely unique to denitrifying bacteria and has been used to indicate the presence of denitri- fiers in the environment. Denitrifiers may belong to organotrophs (organic compounds as energy source), lithotrophs (inorganic compounds as energy source), or (light as energy source). However, the predominant denitrifying bacteria are chemoheterotrophs (Pseudomonas, Aquaspirillum, Azospirillum,andAlcaligenes), photolithoautotrophs (Rhodobacter)orchemo- (Beggiatoa, Thiobacillus, Thioploca,andParacoccus). In coastal sediments, sulfide-dependent denitrification resulting in oxidation of sulfide with nitrate is performed by members of Alpha Beta, Gamma, and Epsilonproteobacteria (Shao et al., 2010). Investigations focused on the occurrence of denitrification, associated processes, and functional diversity of denitrifiers in coastal sediments showed that the major denitrifying communities belong to uncultured

Figure 1.7 Functions of denitrifying microbial communities in sediments (Capone, 1991; Megonigal et al., 2004). Source and composition of organic matter and its role 25 microorganisms clustering within Proteobacteria (Fernandes et al., 2012). TheGammaproteobacteriaarealsoactivemediatorsoftheN,S,andCcycles. About 96% of cultured denitrifiers belongs to the Gammaproteobacteria (Brettar et al., 2002). Nitrate is often limiting at the marine sites of estuaries although hypernutrified estuaries show a high turnover of nitrate where it is constantly resupplied, for example, the Colne estuary shows strong gradients of nitrate and ammonium from the estuary head to the mouth where 20%25% of the total N load entering the estuary is removed by denitrification (Dong et al., 2000). Nitrate reduction is mediated by a diverse polyphyletic group of bacteria (Zumft, 1997) thought to be due to the fact that each bacterial species may participate in only one step of the denitrification process (Burgin and Hamilton, 2007). Wang et al. (2007) showed that addition of electron donors (glucose, sucrose, potato starch, and sodium acetate) stimulated denitrification in Lake Taihu (China) sediments.

1.7.1.3 Iron and manganese reducers The reduction of iron(III) and manganese (IV) by microbes is environ- mentally significant in a variety of aquatic sediments where they have been estimated to oxidize anywhere between 10% and 100% of the OM in aquatic sediments and submerged soils (Lovley, 2006). Most of the cycling of manganese and iron occurs in the oxicanoxic interface of the sediments (Canfield et al., 1993). Microbes that utilize dissimilatory iron (III) and manganese(IV) as TEAs and conserve energy and grow by oxi- dizing organic compounds are phylogenetically and morphologically diverse. These include Geobacteraceae of the Deltaproteobacteria, Geothrix of the Acidobacterium and the Gammaproteobacteria, Ferrimonas, Aeromonas, and Shewanella. The Gammaproteobacteria involved in iron(III) and manga- nese(IV) reduction incompletely oxidize small organic acids and can use a variety of electron acceptors such as nitrate and fumarate (Lovley, 2006). In sediments, iron-reducing microbial communities compete with sulfate redu- cers or methanogens, subject to the limitations of electron donor availability (Fig. 1.8). Most microbes that can reduce iron(III) can also reduce manganese (IV) although in general manganese oxides are found in relatively low concentrations in marine sediments when compared to ferric iron and sul- fate, but like iron, reduced manganese can be reoxidized when diffusing upward to the oxic zone of the sediment. Thus iron(III) and manganese (IV) reduction is generally not the predominant terminal electron accept- ing process in marine estuarine sediments because of the low abundance 26 Microbial Communities in Coastal Sediments

Figure 1.8 Functions of iron- and manganese-reducing microbial communities in sediments. of the amorphous iron and manganese oxide minerals that are readily uti- lized by bacteria. Hence, the complex communities involved in dissimila- tive iron(III) and manganese(IV) reduction during OM degradation in coastal sediments are not much studied compared to the other groups such as fermentative bacteria, nitrate reducers, sulfate reducers, and methanogens.

1.7.1.4 Sulfate reducers Terminal electron accepting processes of sulfate reduction during OM degradation in marine environments are well studied because sulfate reduction dominates in the marine sediments and accounts for up to 50% of OM degradation in marine sediments. SRB are the earliest inhabitants of the earth and sulfate reduction is predicted to evolved approximately 3.7 Ga ago before the evolution of oxygenic (Shen et al., 2001). The diverse physiological abilities of SRB allow them to occur throughout the sediment, while their absolute abundance is limited by the quantity and quality of substrate available. In coastal habitats, the most commonly found SRB belong to the Deltaproteobacteria, with members of the Desulfobacteraceae and Desulfobulbaceae being reported as the most dominant SRB. Anoxic environments are known to be dominated by Deltaproteobacteria (Schwarz et al., 2007). This class of bacteria have Source and composition of organic matter and its role 27 been reported to occur in coastal (Paissé et al., 2008; Zhang et al., 2008), continental shelf (Hunter et al., 2006), as well as cold-seep sediments (Reed et al., 2009). Bacteria belonging to the order Desulfobacterales have been implicated to be involved in sulfur cycling (Vrionis et al., 2005) in particular sulfate reduction (Reed et al., 2009). Members of phylum Deltaproteobacteria not only participate in sulfur cycling but also prevent accumulation of metals (Attri et al., 2011) and inorganic nitrogenous compounds by obtaining energy from the reduction of Fe(III), Mn(IV), and nitrate (Greene et al., 2009). Clostridia members of the Firmicutes and some Euryarchaeota and Crenarchaeota (of the class Thermoprotei) are also known to be sulfate reducers (Muyzer and Stams, 2008). A range of electron donors are found in dissimilatory sulfate reduction metabolism with the commonly cited being acetate and hydrogen. The high of SRB in the clone libraries can be explained by their diverse metabolic capabilities (Devereux et al., 1992; Mussmann et al., 2005; Rabus et al., 2000). They can use a variety of carbon sources, such as different volatile or long-chain fatty acids, alcohols, or aromatic compounds (Coleman et al., 1993; Lovley and Phillips, 1994; Widdel and Hansen, 1992). 22 The process of sulfate (SO4 ) reduction requires eight electrons and 22 SRB invests one ATP to activate SO4 ,whichischemicallystable (Fig. 1.9). The process is mediated by ATP sulfurylase and produces adeno- 22 sine phosphosulfate. This is further reduced to sulfite (SO3 )byAPSreduc- 22 tase, releasing AMP. The SO3 is further reduced to hydrogen sulfide (H2S) by sulfite reductase. The sulfate-reducing reactions and their free energy yield are shown in Table 1.2. The diversity of fatty acids utilized by SRB is reflected in the low con- centrations of all the VFA measured in these sediments, indicating rapid turnover of VFA. The high diversity of SRB throughout the sediment depth is a result of niche partitioning, which is the way in which the dif- ferent species distribute the available resources in the environment. Single-cell studies targeting Dehalococcoides hint at a versatile metabolism, capable of fatty acid and organic compound oxidation as well as carbon dioxide fixation, sulfur cycling, and reductive dehalogenation (Kaster et al., 2014; Wasmund et al., 2014, 2016; Fullerton and Moyer, 2016). Sulfate availability and hence salinity have been reported to be the pri- mary factor determining the relative importance of methanogenesis versus 28 Microbial Communities in Coastal Sediments

Figure 1.9 Functions of sulfate-reducing microbial communities in sediments.

Table 1.2 Free energy yield of sulfate-reducing reactions. Sulfate-reducing reactions ΔG° (kJ/ reaction) 1 22 1 1- 2 1 2 4H2 SO4 H HS 4H2O 151.9 1 22- 2 1 2 2 Acetate SO4 2HCO3 HS 47.6 1 : 22- 2 1 2 1 : 2 1 : 1 2 Propionate 0 75SO4 Acetate HCO3 0 75HS 0 24H 37.7 1 : 22- 2 1 : 2 1 : 1 2 Butyrate 0 5SO4 2Acetate 0 5HS 0 5H 27.8 1 : 22- 2 1 2 1 : 2 2 Lactate 0 5SO4 Acetate HCO3 0 5HS 80.2 sulfate reduction in OC mineralization in estuarine sediments (Middelburg et al., 1993). Mostly under saline conditions, OM mineralization occurs through oxidation pathways ( . 97%) and methane fluxes can be neglected in carbon budgets. Epsilonproteobacteria is metabolically diverse and can uti- lize various electron acceptors including oxygen, nitrate, and different sulfur species (Lin et al., 2006, Nakagawa et al., 2005). Epsilonproteobacteria may live in deep as obligate or facultative chemoorganotrophs by fermenting OM (Lin et al., 2006). They can act as sulfur reducers and/or Source and composition of organic matter and its role 29 oxidizers. The restriction of alpha- and epsilonproteobacteria to the top 2 cm of the sediment may therefore be explained by the high sulfate reduction rates in the top 2 cm of the sediment. Another possible explanation for the detection of clones from alpha-, beta-, and epsilonproteobacteria only in the surface 2 cm of the sediment is mixing with the bottom water because of sediment resuspension. Proteobacteria are most abundant phylum in mangroves and marine sediments. The proteobacterial community in mangroves is dominated by members of the class Deltaproteobacteria, mainly members of the order Desulfobacterales. Dos Santos et al. (2011) reported higher occurrence of Deltaproteobacteria in pristine mangrove sediments. Edmonds et al. (2008) used functional gene clone libraries for SRB (dsrA) to investigate the effect of increased high-molecular-weight carbon on coastal tidal creek sediments. Deltaproteobacteria were found to be the most abundant group in the sulfate-rich zone of White Oak River estuary, North Carolina and harbored the reducing-type dsr genes as well as com- plete sulfate reduction pathways (Baker et al., 2015). However, identifica- tion of nontraditional SRB such as Citrobacter freundii and Bacillus tequilensis in a tropical coastal sediment indicates that the SRB communities are extremely diverse and exhibit geographic patchiness (Vincent and Raj, 2018). Baker et al. (2015) assembled a genomic dataset and identified the metabolic capabilities of various groups of estuarine sediment bacterial com- munities in carbon, iron, nitrogen, and sulfur cycling by metagenomic recon- struction (Fig. 1.10). The identified groups included Betaproteobacteria,

Figure 1.10 Complex interactions between organic carbon utilization, fermentation, and respiration in the bacterial genomes (Baker et al., 2015). 30 Microbial Communities in Coastal Sediments

Gammaproteobacteria, Deltaproteobacteria, Chloroflexi, Planctomycetes, Bacteroidetes, Gemmatimonadetes, Nitrospira, Chlamydiae, and Spirochaetes were similar to the sequences obtained from other estuaries such as Pearl River, Yangtze, and the marine sediments of South China Sea. The genomic analysis exhibits the existence of a high degree of functional redundancy among different groups and highlights the metabolic diversity that promotes essential ecological interactions among the sediment microbial communities. Functional redundancy is a characteristic feature of microbial communities, in which multiple, coexisting but taxonomically distinct organisms can perform similar metabolic functions. Thus it is conceivable that the carbon and nutri- ents cascade through several layers of microbial processing, assimilation as well as remineralization in the sediments. SRB belonging to genera Desulfovibrio and Desulfosporosinus were found to be associated with bioremediation of uranium in contaminated sedi- ments (Cardenas et al., 2008). Proteobacteria are more versatile in their metabolism and can utilize diverse electron acceptors such as organoha- lides. Three species of Deltaproteobacteria active in marine sediments Halodesulfovibrio marinisediminis, Desulfuromusa kysingii, and Desulfovibrio bizertensis were found to reductively dehalogenate bromophenol, by uti- lizing these compounds as TEAs in a process termed as “organohalide res- piration or halorespiration” (Liu and Häggblom, 2018).

1.8 Diversity of archaeal communities Archaea are frequently found within extreme environments, and adapta- tion to chronic energy stress has even been proposed as the primary attri- bute distinguishing them from bacteria (Valentine, 2007). However, archaea are now gaining importance in the global carbon cycle, particu- larly in the coastal sediments. For archaea, many sequences tend to cluster within uncultivated lineages belonging to several dominant phyla (Teske and Sorensen, 2008). Among the archaeal groups, common lineages include Marine Group I (MG-1) (DeLong, 1992), now classified within the phylum Thaumarchaeota (Brochier-Armanet et al., 2008, Zhang et al., 2015), as well as the Miscellaneous Crenarchaeotal group (MCG) (Inagaki et al., 2003b), which has been recently separated into its own phylum, the Bathyarchaeota (Meng et al., 2014). Marine Group II (MG II) of archaea mostly comprise Euryarchaeota (Zhang et al., 2015), which are important players in marine carbon cycle. Advanced identification techniques involving high-throughput sequencing of bacterial community Source and composition of organic matter and its role 31 composition revealed the presence of MG III and MG IV that are lineages of MG II and closely related to MG II in the marine environment. The distribution and abundance of these lineages indicate that Euryarchaeota occupy diverse ecological niches (Fig. 1.11). MG I function as chemolithotrophs and present mostly in the deeper sediments, whereas MG II are and occupy the shallow sedi- ments. Bathyarchaeota are arguably the most widespread and abundant members of sediment archaeal communities, with relative abundances anywhere between 1% and 100% and a very large phylogenetic diversity (Kubo et al., 2012; Fillol et al., 2016). Stable isotopic analyses of subsur- face sediments off Peru have indicated that Bathyarchaeota may have a heterotrophic metabolism, growing by assimilation of fossil OM (Biddle et al., 2006). The MCG group are highly diverse phylogenetically and in terms of their distribution being found in marine and continental habitats (Kubo et al., 2012), but they remain uncultivated so specific physiology is yet to be elucidated, although it is known that most members are hetero- trophic (Biddle et al., 2006). In addition, members of the Marine Benthic Group B/Deep Sea Archaeal Group (MBG-B/DSAG) are also ubiquitous in subsurface envir- onments (Vetriani et al., 1999; Inagaki et al., 2003a). This lineage has recently been classified into its own separate candidate phylum, the Lokiarchaeota (Spang et al., 2015) within the Asgard superphylum

Figure 1.11 Global pattern of Archaeal distribution (Lloyd et al., 2013). 32 Microbial Communities in Coastal Sediments

(Zaremba-Niedzwiedzka et al., 2017). Additional approaches using single- cell genomics and metagenomics have suggested a role in the degradation of detrital proteins as well as acetogenesis (Lloyd et al., 2013; Lazar et al., 2016). Archaeal community is dominated by the Crenarchaeota in the sulfate reduction phase and Euryarchaeota in the methanogenic phase. Methanomicrobia are also dominant group that increase in relative abun- dance with depth. This class includes members of the anaerobic methano- trophic archaea that are often enriched in aggregates with sulfate reducers in the SMT zone (Boetius et al., 2000).

1.8.1 Methanogenic archaea Methanogenesis is a unique terminal energyyielding process exclusive to methanogenic archaea (MA) that occurs via two main pathways: CO2 reduction or the formation of methane from methylated compounds (MA displays heterotrophic metabolism) and autotrophy via the acetyl-Co-A pathway. Methanogens have the ability to use only a limited number of substrates, including H2,CO2, formate, acetate, methanol, and methylated amines (Oremland, 1988). The most important substrates for methanogens are H2/CO2 and acetate, and they often depend on other anaerobic fer- mentative bacteria for these substrates (Conrad, 1999). The hydrogeno- trophic MA (methanogenic species use H2/CO2 as substrate) outnumber the acetoclastic MA (the methanogens that use acetate as substrate) (Garcia et al., 2000). MA belong to the phylum Euryarchaeota that are classified into seven orders (Methanosarcinales, Methanobacteriales, Methanomicrobiales, Methanocellales, Methanococcales, Methanomassisilicoccales, and Methanopyrales). However, the early hypothesis that methane metabolism originated early in the evolution of Euryarchaeota has been challenged following the discovery of methane metabolism by members of other phyla including Miscellaneous Crenarchaeota Group (MGB; presently Bathyarchaeota) (Evans et al., 2015). Methanogenesis performed by MA is catalyzed by methyl-coenzyme M reductase (MCR). The mcrA gene encoding a subunit of enzyme MCR is a commonly used gene marker in molecular surveys (Conrad, 2007; Bridgham et al., 2013). The advan- tage of the mcrA gene marker is to capture both the phylogenetic and functional signature of methanogens (Borrel et al., 2013). However, the of acetoclastic versus hydrogenotrophic methanogenesis has been reported to be related to sediment properties (i.e., pH and temperature). Phelphs and Zeikus (1984) reported that Source and composition of organic matter and its role 33

acetoclastic methanogenesis was the major pathway for CH4 production in mildly acidic (pH 6.2) conditions. The increase in pH to neutral values enhanced total CH4 production from H2/CO2, but did not affect the CH4 produced from acetate. Other studies have shown that CH4 produc- tion at low temperature (4°C) in sediments was mainly from acetate; however, an increase in temperature (20°C25°C) leads to an increase in contribution of CH4 production from H2/CO2 (Schulz and Conrad, 1996; Schulz et al., 1997; Nüsslein and Conrad, 2000; Glissmann et al., 2004). Thus methanogenic community structure is thought to be related to the availability of “competitive substrates,” for example, Purdy et al. (2002) compared a predominantly freshwater sediment site of the Cole estuary with a marine site with phytotypes closely related to the specialist obligate acetate-utilizing Methanosaeta concii found at the sulfate-limited freshwater site, and the more generalist Methanogenium being found at both sites of the estuary. Methanogens commonly detected in marine, salt-marsh, river, estuarine, and tidal flat sediments belong to Methanosarcinales and Methanomicrobiales. Parkes et al. (2012) found a close relationship between the depth distribution of methanogenic substrate utilization and methanogens that can utilize these compounds. Methylamine-utilizing Methanosarcinales were dominant in the near-surface sulfate reduction sediments and Methanosaeta were dominant in the deeper sediment layers. Noncompetitive substrates such as methanol and trimethylamine utilized by many Methanosarcinales sp. are important substrates for methanogenesis especially in sulfate-containing anoxic sediments (Oremland and Polcin, 1982). The distribution and prevalence of MA are also thought to be largely restricted by the presence of other alternative electron acceptors such as sulfate that allow the SRB to grow more thermodynamically. Sulfate- reducing anaerobic bacteria can outcompete methanogens for H2/CO2 and acetate due to higher substrate affinities and higher energy and growth yields (Lovley and Phillips, 1986); however, both processes can coexist (Wand et al., 2006). Coexistence occurs because of spatial varia- tion in the abundance of TEAs or because the supply of electron donors is nonlimiting (Roy et al., 1997; Megonigal et al., 2004). Environmental factors also have an important role in designing the methanogenic archaeal communities in coastal sediments. For example, in low-salinity sediments, the dominant species of methanogens were Methanosaeta and putatively hydrogenotrophic Methanomicrobiales, whereas, the marine sediments were dominated by methylotrophic Methanococcoides and versatile Methanosarcina (Webster et al., 2015). Generally, it has also 34 Microbial Communities in Coastal Sediments been reported that increasing salinity inhibits hydrogenotrophic methanogens but enhances acetoclastic methanogenesis (Liu et al., 2019). Methanosarcina has ahighgrowthratebutalowaffinity for acetate, while Methanosaeta has a low growth rate but a high affinity for acetate ( Jetten et al., 1992). Methanosaeta spp. are generally distinguished from other Methanosarcinales by their obligate requirement for acetate, which is used as a sole substrate for methanogenesis and also differentiated based on their rod-shaped cells (Ma et al., 2006). Hence, there is an interesting possibility of among methylotrophic methanogens due to competition for C1 substrates (Antony et al., 2012). The prevalence of methylotrophic methanogens in the marine- dominated zone reveals that methylotrophic methanogenesis is favored over hydrogenotrophic and acetoclastic methanogenesis in sulfate-containing sedi- ments because of competition from SRB for hydrogen and/or acetate (Oremland and Polcin, 1982). Methylotrophic methanogenesis has been shown to be an important biogeochemical component of carbon cycling in the Big Soda Lake, USA (Oremland et al., 1982). Anthropogenic activities have shown to increase the potential produc- tion of CH4 in pristine systems (Reshmi et al., 2015; Sanders et al., 2007). Saia et al. (2010) reported the presence of methanogens in tropical estua- rine sediments that are highly contaminated, especially with pollutants of petroleum source. Hence, understanding these potential changes in situ and how the microbial community may respond to reoxidation events in relation to the efficiency of OM degradation, nutrient cycles, and terminal output processes requires further investigation.

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2.1 Introduction The earth is a closed system and hence, the amount of nutrients (N and P) in all forms is essentially fixed. Nevertheless, these nutrients are constantly recycled and combine with different elements to form variety of chemical compounds (Boesch, 2002). Human activities have resulted in large-scale changes in the during the last two centuries, with a pro- nounced increase since the 1950s (Galloway et al., 2003, 2008). Each ele- ment has its own physicochemical and biogeochemical properties and also a unique chemistry within air, water, and land (Boesch, 2002). Human activi- ties have profoundly influenced the global cycling of nutrients, especially movement of nutrients to estuaries and other coastal waters (Howarth et al., 2000). Due to an increase in demand for fuel and fiber, the global annual nitrogen input to ecosystems also increased (Smil, 2002). In addition, the global nitrogen rise in input is expected to continue and projections for 2030 show a significant increase (Seitzinger et al., 2010). Coastal waters account for 10%15% of total sea area globally and they support nearly 50% of marine primary production (Paerl, 1997). The rapid increase of human activities has increased nutrient transport from land to sea in the past decades, resulting in environmental deterioration as well as changes to biogeochemical processes (e.g., Humborg et al., 1997; Rabalais et al., 2000; Qu and Kroeze, 2010; Seitzinger et al., 2010). For example, human activities have more than doubled the global availability of N and P to biological processes. Although anthropogenic sources of nitrogen and phosphorus to coastal waters are surprisingly similar in different marine areas of the world, human-driven changes in nutrient cycling have not occurred uniformly around the world (Howarth et al., 2000). The key factors behind the few observed differences in nutrient sources are population density, land use, traffic, use of artificial fertilizers, industrial emissions and other combus- tions, and progress in sewage management (Vitousek et al., 1997).

Microbial Communities in Coastal Sediments © 2021 Elsevier Inc. DOI: https://doi.org/10.1016/B978-0-12-815165-5.00002-9 All rights reserved. 47 48 Microbial Communities in Coastal Sediments

Moreover, it cannot be denied that such changes are pronounced and con- centrated in the areas of high human population density and intensive agri- cultural production (Howarth et al., 2000; Seitzinger et al., 2010). These conditions are particularly obvious in tropical coastal zones that receive the lion’s share of inputs of both freshwater and sediments from rivers (Fig. 2.1). The first-order controls of substances transported into the coastal environ- ment by rivers are natural factors, whereas, the second-order control are human activities, which gained importance in the “Anthropocene,” aperiod starting with the beginning of industrialization in the late 18th century, and in particular during the second half of the 20th century (Jennerjahn, 2012). Increased river loadings of N, P, and organic matter are nowadays among the major threats to human water security and river water quality and biodi- versity (Vörösmarty et al., 2010). Economic development has increased nutrient loads and influenced nutrient cycling in coastal regions, resulting in accelerated degeneration of the coastal ecosystem and the occurrence of eutrophication. In developed nations in Europe, North America, Asia, and Oceania, extreme nutrient enrichment took place mainly over the narrow period of 196080. This 20-year period coincided with a five-fold increase in the use of manufac- tured fertilizers. It also led to a rapid growth in scientific investigations as well as in publications on coastal nutrient enrichment (Boesch, 2002; Nixon, 2009). However, most of the developing world lies in the tropics or subtropics and it is the coastal marine ecosystems in these regions that will be most severely impacted by nutrient pollution in the coming dec- ades (Nixon and Fulweiler, 2009). Tropical coastal ecosystems are the most biogeochemically active zones and are more easily affected by anthropogenic nutrient loading than those in higher latitudes (Yule et al., 2010; Smith et al., 2012). This is particularly the case in the tropical S and SE Asia rivers and coastal regions because of their high natural inputs in combination with extensively modified catchments and resulting degrada- tion as a consequence of rapid economic development and population growth (Jennerjahn et al., 2004, 2008; Borbor-Cordova et al., 2006; Pan et al., 2007; Jennerjahn, 2012; David et al., 2016). The developing world is also the place where human population growth is greatest, the most rapid growth is in urban areas, particularly near the coast, which will bring increasing amounts of nutrients to the coast. These systems are highly sensitive to natural environmental changes and human activities, including land use, geomorphologic evolution, and deforestation. In spite of this, the nutrient characteristics and the Sources, types, and effects of nutrients (N and P) in coastal sediments 49

(A)

,

Total =36,000 km3 year-1 (B)

Total = 19,000 * 106 t year-1 Figure 2.1 Global river fluxes of (A) water and (B) sediment into the ocean. Major drainage regions are denoted by different colours. The direction and strength of arrows represents the direction and annual flux of water and sediment input. The numbers denote discharge in km3 year21 for water (A) and in 106 t year21 for sedi- ment (B). White areas on land are regions which have no discharge of water and sed- iment into the ocean. From Milliman and Farnsworth (2011). processing of nutrients in tropical rivers and estuaries are poorly studied relative to those in temperate zones (Jennerjahn, 2012). This chapter ana- lyzes and compares the various sources of nutrients, mainly N and P, to coastal ecosystems and their effects on eutrophication that is the major 50 Microbial Communities in Coastal Sediments

Figure 2.2 Global overview of eutrophic and hypoxic coastal areas, as well as coastal systems in recovery. Taken from World Resources Institute (www.wri.org), Diaz and Selman, 2010. consequence of nutrient enrichment (Fig. 2.2). Further, eutrophication- induced sediment hypoxia and resultant changes in microbial community structure in the coastal sediments are discussed.

2.2 Nutrient sources of coastal ecosystems The major natural sources of reactive nitrogen on earth are the weather- ing of lithospheric rocks and biological fixation of dinitrogen gas from the atmosphere (Vance, 2001; Holloway and Dahlgren, 2002). The major natural source of phosphorus is the weathering of rocks (Berner and Berner, 2012). Apart from natural sources, nutrients enter the coastal envi- ronment due to two essential human activities: (1) the combustion of organic matter to release energy (including biomass, coal, oil, and natural gas) and (2) the production and consumption of food (Galloway et al., 2002). These land-based activities result in agricultural runoff, sewage and industrial discharges, and fossil fuel combustion, and atmospheric deposi- tion results in nutrient enrichment in the coastal marine environment (GESAMP, 2001). Among these sources, sewage and industrial discharges Sources, types, and effects of nutrients (N and P) in coastal sediments 51 are point sources, whereas agricultural runoff and atmospheric deposition are nonpoint sources. Population density is a major factor that determines the point source of nutrient input into coastal ecosystems. Nutrient feed- ing from nonpoint source is particularly high when a watershed is heavily populated. Smith et al. (1999) used data sets from 165 rivers that resulted from LOICZ (Land-Ocean Interactions in the Coastal Zone), a core proj- ect of the International Geosphere-Biosphere Programme (IGBP), to assess the global nitrogen and phosphorus inputs in the ocean and the controlling factors. From the statistical evaluation of these data and rela- tions to potential drivers, they found that dissolved inorganic nitrogen and dissolved inorganic phosphorus yields can be generally parameterized by population density and runoff per area on a global scale. The total loads for the 1990s are about three times higher than in the 1970s and the high- est yields can be found in tropical S and SE Asia (Smith et al., 1999). However, in coastal ecosystems that have less populated catchments, nutrient inputs are high from point sources than nonpoint sources. The relative size of a catchment to that of the coastal ecosystem also discrimi- nates the importance of point or nonpoint sources of nutrient input into the coastal ecosystem. For example, in an estuary, when the catchment is heavily populated and is smaller than the size of the estuary, point source can be a major contributor of N to the estuary (Howarth et al., 2000). Although agriculture is the largest activity controlling nonpoint fluxes of phosphorus, for nitrogen, both agriculture and fossil fuel combustion con- tribute to nonpoint source flows to coastal ecosystems. In most estuaries, N and P inputs from nonpoint sources are greater than those from waste- water. This is pronounced in estuaries that have large watersheds with more area to trap N from atmosphere and also more land devoted to crop and livestock production. For example, the contribution of N and P by point sources such as wastewater treatment plants represents only one- quarter of the total inputs into Chesapeake Bay (Howarth et al., 1996). In coastal zones, the primary source of inorganic nutrients is water- sheds, whereas, that of organic matter is watersheds and tidal wetlands (Hopkinson et al., 2019). Land use, land cover, and land management practices are the primary factors controlling long-term variability in organic matter and nutrient export to coastal ecosystems. The relative bal- ance between organic matter and inorganic matter loading determines the net metabolism as well as the overall balance between the gross primary production and respiration in coastal ecosystems. The degree to which microbial decomposition of allochthonous organic matter is stimulated in 52 Microbial Communities in Coastal Sediments estuarine systems is a function of the relative balance between inorganic/ organic turnover times and water residence time in the estuary. In condi- tions where the ratio of inorganic and organic loading is constant, a decreasing residence time transforms the system toward increased autot- rophy, while increasing residence time pushes the system toward hetero- trophy (Hopkinson and Vallino, 1995).

2.2.1 Agriculture One of the most pervasive causes of coastal ecosystem degradation is increased supply of nutrients from agriculture. Synthetic inorganic fertili- zers invented during World War I came into a widespread use during the late 1950s. The rate of fertilizer use increased steadily until the late 1980s, when the collapse of the former Soviet Union led to great disruptions in agriculture, which led to a slight decline in global N fertilizer use for a few years, which regained after 1995 (Figs. 2.3 and 2.4; Howarth et al., 2000). Approximately half of the inorganic N fertilizer that has ever been used on earth has been applied during the past few decades. Moreover, inorganic fertilizers account for more than half of the human alteration of the nitrogen cycle (Vitousek et al., 1997). Other human controlled pro- cesses such as production of nitrogen-fixing crops in agriculture also con- tribute to the problem (Boesch, 2002). The amount of nitrogen and 2 phosphorus fertilizer produced was 109 Tg N year 1 and 47 Tg 21 P2O5 year , respectively, in 2014 (Roser and Richie, 2020). The global production of nitrogen fertilizers was ,10 million metric tonnes in 1950, but reached 109 million metric tonnes in 2014 and is projected to exceed 135 million metric tonnes of N by the year 2030 (Vitousek et al., 1997; FAO, 2011; Roser and Richie, 2020). The demand for phosphate was 47 million metric tonnes in 2011 with an expected annual growth rate of 1.9% (FAO, 2011; Roser and Richie, 2020). Two factors have been iden- tified as major drivers for the ever-increasing use of synthetic fertilizers: increasing human population and a growing world economy (Steffen et al., 2007). East Asia is the largest of fertilizers in the world (37% of all consumption), and Asia as a whole consumes 60% of the world’s demand. The consumption in Europe (13%), North America (13%), South America (10%), and Oceania (1.6%) is comparatively smaller (Korpinen and Bonsdorff, 2015). For example, in one of the largest rivers of the world, the Mississippi, human activities have increased nitrogen influx four-fold (UNEP/GPA, 2006). The intensive application of Sources, types, and effects of nutrients (N and P) in coastal sediments 53

Figure 2.3 Spatial distribution of global agricultural nitrogen (N) fertilizer use in the year 1961, 1980, 1990, and 2013. Colors show N fertilizer use rate in per square meter cropland of each pixel. Taken from Lu and Tian (2017). agrochemicals in the river catchment led to a drastic increase in the dis- solved nitrogen load of the river. Attendant with an increase of nitrogen 2 fertilizer use in the United States from about 3 3 106 t year 1 in the 2 1960s to 12 3 106 t year 1 in the 1990s, the nitrate concentration in the river increased from about 40 to 120 μM(Goolsby et al., 2000). Overall agricultural production has more than doubled, yet production relies on less land and on fewer but larger farms. For P, agriculture is one of the largest sources of nonpoint pollution, whereas for N, both agriculture and burning of fossil fuels contribute significantly to nonpoint flows to estuar- ies and coastal waters. Hence, the primary cause of coastal eutrophication worldwide is attributed to the rapid intensification of agriculture (Matson et al., 1997). Surplus N and P from agricultural fertilizers accumulate in soils and partly move into surface waters via land erosion, surface runoff, and melt waters. The total amount of nutrients exported in runoff from the landscape to surface waters increases linearly with the soil nutrient content. N migrates into groundwater or enters the atmosphere via ammonia volatilization and 54 Microbial Communities in Coastal Sediments

Figure 2.4 Spatial distribution of global agricultural phosphorus (P) fertilizer use in the years 1961, 1980, 1990, and 2013. Colors show P fertilizer use rate in per square meter of cropland in each pixel. Taken from Lu and Tian (2017). denitrification. N from these sources can reach the water either by direct leaching or runoff from farm fields or indirectly from the atmosphere. N in theatmosphereisbyvolatilizationofammonia that is lost from the inorganic N fertilizer that is applied to agricultural crops and animal wastes and lost to the air. For example, in the United States, the amount of ammonia volatil- ized to the atmosphere from agricultural systems is roughly equivalent to the amount of nitrate that leaches from crop fields into surface waters. Since ani- mal wastes are a major source of airborne N in addition to the nutrients that animal wastes leak directly to surface waters, animal wastes may be the largest single source of N from agricultural operations in coastal waters (Howarth et al., 2000).

2.2.2 Animal husbandry and marine aquaculture Animal waste may also be a single largest source of N from agricultural production to coastal waters, either directly through runoff or indirectly Sources, types, and effects of nutrients (N and P) in coastal sediments 55 through volatilization and deposition of atmospheric nitrogen (Boesch, 2002). Prior to World War II, farming communities (animal husbandry and agriculture) were self-sufficient, in the sense that feed produced locally from agriculture could meet animal requirements, and manure from the livestock could be recycled to meet crop fertilization needs (Howarth et al., 2000). Increased global meat consumption by 3% per year since the 1960s (FAO, 2006), as well as increased fertilizer production and availability encouraged the decoupling of the grain and animal pro- duction systems. The consequent industrial-scale production and separa- tion of crop and animal production eventually led to import .80% of grain for feed by the major livestock-producing countries. This evolution has resulted in accumulation of N and P in soils of animal producing areas. Moreover, this excess of nutrient-laden manure has tended to build up in less productive agricultural areas with a limited capacity to make use of the nutrients in manure in crop production. This has exacerbated pro- blems of the nonpoint runoff of manure nutrients into watersheds and ultimately, coastal waters. Animal waste also contributes to gaseous nitrogen emissions, which when deposited back onto the landscape, can also be a significant source of N in coastal marine ecosystems. Meat production is very inefficient in terms of N and the consumption of meat protein is of particular concern in terms of N pollution (Nixon and Fulweiler, 2009). It requires 100 kg of N in corn (maize) to produce 5 kg of edible N in beef, with a protein conversion efficiency of only 5%. The remaining 95 kg of N ultimately enters the landscape as metabolic wastes from the animals or their car- casses. Increased consumption also leads to excretion of N into the envi- ronment, which varies with the type of protein consumed, vegetable or meat protein. Aquaculture and fish farming are important and large point sources of direct nutrient input into coastal waters. Global fish production has reached 171 million tonnes per year in 2016, 47% of which were from aquaculture (FAO, 2018). Many aquaculture operations invariably result in the release of metabolic waste products (feces, pseudo-feces, and excreta) including ammonia and urea and uneaten food into the aquatic environment. The largest proportion of the waste, predominantly organic carbon and nitrogen, settles in the sediment in the immediate vicinity of the farm. The release of soluble inorganic nutrients (N and P) has the potential to cause eutrophication in aquaculture facilities themselves, as well as in adjacent coastal waters. Fish farms can affect coastal ecosystems 56 Microbial Communities in Coastal Sediments by a constant, year-round input of nutrients and organic matter, which can prolong the season for primary production and eliminates the natural nutrient limitation. The fish farms in the coastal and archipelago areas of the Baltic Sea, for example, contribute 15 times the load from municipal wastewater for P and 36 times the nitrogen input through treated wastewater (Bonsdorff et al., 1997). A global assessment of the nutrient input from marine aquaculture into coastal waters based on modeling and FAO statistics concluded that it is a quantitatively significant source. It was also projected to increase six- fold until 2050, with China being the quantitatively most important and most strongly increasing source (Bouwman et al., 2013). Extensive studies on the Chinese island of Hainan have demonstrated the deleterious effects of not well-managed intensive brackish pond aquaculture on the well- being of adjacent coastal ecosystems (Jennerjahn et al., 2013; Herbeck et al., 2013; Roder et al., 2013). Large parts of the NE coast of Hainan are covered by aquaculture ponds in which mainly fish and shrimps are grown and harvested once per year in the case of fish and three to four times per year in the case of shrimps. The untreated effluents from the ponds are released directly into coastal waters that are harboring seagrass meadows and coral reefs. Since the 1960s tens of square kilometers of ponds to a large extent replaced coastal mangrove forests, that is, a sink for land-derived nutrients has been replaced by a source (Herbeck et al., 2 2020). A total of 391 3 106 m3 year 1 of aquaculture effluents are being released from a pond area of 40 km2 every year, which results in a total of 1292 tonnes of nitrogen, half each in the dissolved and particulate forms, of 51 tonnes of phosphorous, and of 6151 tonnes of organic carbon (Herbeck et al., 2013).

2.2.3 Fossil fuel burning and atmospheric deposition The by-products of fossil fuel combustion principally, exhaust from motor vehicles (mobile source) and electric power generation (fixed source), are major sources of N to coastal waters in many regions. Nevertheless, the relative importance of these combustion sources of reactive N varies around the world. For example, in Asia, road transport accounted for nearly 28% of N oxide emissions in 1990 that was 45% in Europe in 1998 (Bradley and Jones, 2002). However, Asia contributes to larger share of N oxide emissions in coal burning than Europe, which relies more on oil, natural gas, etc. for electric power generation. Nitrogen-based trace gases Sources, types, and effects of nutrients (N and P) in coastal sediments 57 such as nitric oxide released during fossil fuel combustion may be directly deposited onto the coastal waters. Additionally, those deposited on land- scape as acid rain or dry pollutants can run off into surface water and thus reach coastal ecosystems. Direct atmospheric deposition of nitrogen may contribute between 1% and 40% of the total nitrogen inputs to coastal ecosystems and is most pronounced near to emission sources (Fig. 2.6). This also depends on the size of the coastal ecosystem relative to its water- shed; the larger the size, the greater will be the percentage of N that is deposited (Korpinen and Bonsdorff, 2015). In contrast to P, the amount of N exported into coastal waters from nonagricultural landscapes, includ- ing forests, can be substantial. In Chesapeake Bay, atmospheric depositions comprise 12% and 6.5% of TN (total nitrogen) and TP (total phosphorus) inputs, respectively (Kemp et al., 2005). With relatively larger watersheds than the size of the coastal ecosystem, the greater source will be run off of N deposited in landscapes. This may be more important than direct depo- sition and more difficult to quantify (Nixon and Fulweiler, 2009). As fuel combustion releases reactive or biologically available N into the atmosphere, it can easily travel great distances before deposited on land and water. Hence, the source of atmospheric N deposited in coastal watersheds can be far from the coast and even outside the watershed area that drains into a bay or estuary. The area from which various materials may be put into the atmosphere and reach a given estuary is called the airshed of that estuary (Fig. 2.5). For example, the airshed of Chesapeake Bay is 6.5 times larger than the watershed of the bay, which is nearly 17 times larger than the bay

Figure 2.5 Source of nutrients and routes of transfer to coastal waters. Figure 2.6 Estimated total reactive nitrogen deposition from the atmosphere (wet and dry) in 1860, early 1990s, and projected for 2050. Created by Philippe Rekacewicz, Emmanuelle Bournay, UNEP/GRID-Arendal (www.grida.org). From: Galloway, J.N., Dentener, F.J., Capone, D.G,. et al., Nitrogen cycles: past, present, and future. Biogeochemistry 70, 153226. Sources, types, and effects of nutrients (N and P) in coastal sediments 59

(Nixon and Fulweiler, 2009). In addition to fossil fuel burning, agriculture contributes to N in the atmosphere in two ways: from cultivation of N-fixing crops as well as from industrial production of inorganic N fertilizers by the HaberBosch process (Smil, 2002). The production of reactive N from agriculture is five times greater than that from fuel combustion (Galloway et al., 2002). On a global scale, atmospheric nitrogen deposition has strongly increased during the Anthropocene with centers mainly in densely populated and intensively used regions (Fig. 2.6).

2.3 Nutrient enrichment: forms and types Among the several environmental problems faced by coastal ecosystems, nutri- ent enrichment is the most significant one. A study conducted to compare global and regional N budgets suggested a future with more than 70% of reac- tive N moving through the biosphere, and nearly 30% increase in reactive N reaching the coast (Galloway et al., 2002). Elevated concentrations of dissolved organic N and dissolved inorganic N (DIN) in coastal regions result primarily from freshwater inputs from the terrestrial environment as well as from atmo- spheric deposition. While most of this N is removed by denitrification in the coastal sediments, biological N2 fixation serves as an important in situ source of N in these systems (Gruber, 2008). Land-based inputs dominated by 1 2 2 human activities provide nitrogen as NH4 ,NO3 ,NO2 ,orinanorganic 1 2 form, whereas, from atmospheric emissions, the forms are NH4 ,NO3 ,or 2 NO2 (Korpinen and Bonsdorff, 2015). While nitrogen in agriculture efflu- ents is usually dominated by nitrate, aquaculture effluents contain mainly ammonium (e.g., Goolsby et al., 2000; Herbeck et al., 2013). Under aerobic 1 2 conditions, NH4 is nitrified to NO3 . However, by vertical mixing as well 1 2 as by upwelling, NH4 and NO3 are available to primary producers in the euphotic zone. The N bound in organisms sinks to the sediment and under anaerobic conditions, denitrification converts nitrite, nitrate, and ammonium to gaseous N2.Inadditiontodenitrification,another process of anaerobic ammonium oxidation (Anammox) also removes N from the coastal environ- ment. Hence, denitrification and Anammox are considered as significant major sinks of N in various coastal ecosystems. Dissimilatory nitrate reduction to ammonium (DNRA) is a direct reduction of nitrate to ammonium and retains 1 1 NasNH4 . Highly reduced sediments retain more NH4 due to denitrifica- tion and also due to elevated levels of DNRA (Middelburg and Levin, 2009; Fig. 2.7). 60 Microbial Communities in Coastal Sediments

Figure 2.7 Conceptual model of N input and transformation in the coastal environment.

In coastal waters, N removal by denitrification is usually greater than inputs from land, which helps to maintain the ecosystem in an oligotro- phic and N-limited state (Hansell and Follows, 2008). However, there are exceptions. For example, certain temperate estuaries like, for example, the Apalachicola on the Gulf coast of Florida and several estuaries on the North Sea coast of the Netherlands are P limited. This P limitation is due to stringent controls on P releases combined with high and unregulated anthropogenic N inputs (Howarth et al., 2000). Although not large, pri- mary producers can have an effect on sediment denitrification rates. In shallow coastal areas, primary producers outcompete denitrifying bacteria for DIN, thus inhibiting denitrification (Dalsgaard et al., 2003). In tropical coastal systems with carbonate sands and little human activ- ity, P is generally limiting to primary production, because the sand readily adsorbs phosphate, trapping it in the sediment and leaving it largely unavailable to organisms. However, such lagoons may move toward N limitation as they become eutrophic. The primary reason is that as more nutrients enter these waters, the rate at which sediments adsorb phosphate slows and a greater proportion of the P remains biologically available (Howarth et al., 2000). With increasing N retention in the sediment, the N:P ratio also decreases. This causes a shift in phytoplankton communities toward cyanobacterial dominance. Sources, types, and effects of nutrients (N and P) in coastal sediments 61

2.4 Effect of hypernutrification 2.4.1 Eutrophication and consequences for ecology Eutrophication is a global anthropogenic pressure on coastal ecosystems. Sheltered coastal systems, such as estuaries and bays, are particularly vulnerable to impacts of eutrophication (Korpinen and Bonsdorff, 2015). The past few decadeshavewitnessedamassiveincreaseineutrophicationglobally,leadingto hypoxic and anoxic conditions in coastal ecosystems (Howarth, 2008). A myr- iad of direct or indirect biogeochemical as well as ecological responses to anthropogenic fertilization of ecosystems are observed (Fig. 2.8; Cloern, 2001). Specifically, nutrient enrichment contributes to autochthonous eutrophication in coastal ecosystems by increasing primary production, thus resulting in increased supply of organic matter within the ecosystem itself. In contrast, allochthonous eutrophication occurs when there is an increasing supply of organic matter into

Figure 2.8 Schematic diagram of the different pathways of nutrient deposition into coastal waters and ensuing processes leading to eutrophication (algal blooms) and hypoxia. Taken from www.coastalwiki.org/wiki/Portal:Eutrophication, Hans W. Paerl. 62 Microbial Communities in Coastal Sediments the coastal ecosystem from outside the system (Nixon and Fulweiler, 2009). The first wave of coastal eutrophication was allochthonous and occurred during the second half of the 19th century due to the input of large amounts of organic matter from industries as well as mixing of sewage laden rivers into coastal ecosystems. However, autochthonous eutrophication emerged as a seri- ous concern much later (Nixon, 1995). In addition to nutrient enrichment as a major cause of autochthonous eutrophication, other reasons like, for example, the construction of dams, which reduces the transport of suspended sediment downstream, are also attributed. The decline in the supply of organic matter can increase the primary production due to an increase in the clarity of water. The major nutrients that cause eutrophication due to over-enrichment are N and P. It is generally conceived that eutrophication in coastal marine ecosystems is controlled by N inputs, whereas, in freshwater lakes, it is by P inputs. Nevertheless, in some coastal ecosystems, the identity of the nutrient that limits eutrophication switches seasonally between N and P. Tidal mix- ing may reduce the effect of eutrophication in spite of higher N and P loading rates, for example, San Francisco Bay receives higher N and P load- ing than Chesapeake Bay, but is not as sensitive to eutrophication as Chesapeake Bay. This is due to intensive dilution and flushing by freshwa- ter inflow and tidal mixing (Cloern, 2001). The residence time of water in a coastal system determines to a large extent the magnitude of the effect that nutrients have on its primary production and its trophic state. For example, if the residence time is less than 1 month approximately 70% 80% of nitrogen and phosphorus will be exported and hence will not con- tribute to eutrophication in the coastal water body (Figs. 2.9 and 2.10). The single largest coastal system affected by eutrophication in the United States is the so-called “dead zone” in the Gulf of Mexico, an extensive area of reduced oxygen levels. In the early 1990s, the zone cov- ered an estimated 9500 km2 of the gulf, extending out from the mouth of the Mississippi River. By the summer of 1999, this hypoxic area had dou- bled to 20,000 km2 (Boesch, 2002; Howarth et al., 2000, Rabalais et al., 2000). Other coastal systems that were severely impacted include Chesapeake Bay, Long Island Sound, and the Florida Keys in the United States; the Baltic, North Adriatic, and Black Seas of Europe are all exam- ples of eutrophication as a result of nutrient over-enrichment (Fig. 2.11). In the Black Sea, eutrophication was partially reversed during the early 1990s as the nutrient inputs decreased following the collapse of the Soviet Union and fertilizer use in Eastern Europe dropped sharply. This decrease Sources, types, and effects of nutrients (N and P) in coastal sediments 63

Figure 2.9 The percent of total nitrogen input from land and atmosphere that is exported from a sample of estuaries and lakes as a function of mean water residence time in the system. The regression line is for the combined data set. Separate regressions for estuarine and lake data were virtually identical. Taken from Nixon, S.W., Ammerman, J.W., Atkinson, L.P., Berounsky, V.M., Billen, G., Boicourt, W.C., et al., 1996. The fate of nitrogen and phosphorus at the land-sea margin of the North Atlantic Ocean. Biogeochemistry 35, 141180. For details of data shown see the original publication. was temporary, however, and both nutrient inputs and eutrophication in the Black Sea have reached an all-time high (Howarth et al., 2000). The general rule of thumb is that phosphorus controls freshwater eutrophication and nitrogen is the controlling factor of eutrophication in coastal marine ecosystems. However, it is important to manage both phos- phorus and nitrogen inputs because in some systems, managing nitrogen without managing phosphorus inputs can lead to a situation where phos- phorus becomes the nutrient controlling eutrophication. In ecosystems such as Chesapeake Bay (Malone et al., 1996) and the Gulf of Mexico including the “dead zone” (Rabalais et al., 1999), nitrogen is probably the nutrient responsible for the impacts of eutrophication. When primary pro- duction in these systems is phosphorus limited, relatively little biomass tends to sink into bottom waters and the biomass that sinks out of the water column is more likely to be controlled by nitrogen than by phos- phorus (Glibert et al., 1995; Malone et al., 1996). 64 Microbial Communities in Coastal Sediments

Figure 2.10 The percent of total phosphorus input from land that is exported from a sample of estuaries and lakes as a function of mean freshwater replacement time in the system. Regression line includes lakes and estuaries but excludes the Guadalupe in a low flow (long residence time) year. Taken from Nixon, S.W., Ammerman, J.W., Atkinson, L. P.,Berounsky,V.M.,Billen,G.,Boicourt,W.C., et al., 1996. The fate of nitrogen and phospho- rus at the land-sea margin of the North Atlantic Ocean. Biogeochemistry 35, 141180. For details of data shown see the original publication.

Figure 2.11 Locations of regions of large-scale nutrient over-enrichment. Taken from Boesch, D.F., 2002. Challenges and opportunities for science in reducing nutrient over- enrichment of coastal ecosystems. Estuaries 25 (4), 886900. Sources, types, and effects of nutrients (N and P) in coastal sediments 65

Eutrophication mainly influences coastal zones and according to UNEP/GPA (2006), it is also apparent over larger areas of semienclosed seas leading to changes in the structure of ecological communities by at least two mechanisms: directly by altered nutrient concentrations compo- sition and indirectly through oxygen depletion related to the decomposi- tion of the high biomass produced. Nutrient over-enrichment alters community structure directly by changing competition among algal spe- cies for nutrients. A prime example in this respect is the NW Black Sea that is fueled by the Danube River, the second largest river of Europe. Major changes in economy and land use in Romania, in which the Danube discharges into the Black Sea, altered the amount and composi- tion of nutrients in coastal waters. In the late 1960s and early 1970s and then in the 1980s the two major “Iron Gates” dams were built in the Danube approximately 800 km from the coast that altered the river flow strongly (Popa, 1993). In combination with an increase in effluents from industrialization and agriculture, it also altered the amount and composi- tion of nutrients of the Danube. The damming of the river reduced tur- bulence and light attenuation in the artificial lake and allowed for extensive primary production, mainly in the form of diatoms, algae with a siliceous frustule. The river water leaving the reservoir was depleted in nitrate, phosphate, and silicate, but was refueled with nitrate and phos- phate during its passage through the lowlands with their agriculture, industrial, and municipal effluents. However, silicate, which almost exclu- sively stems from natural weathering, remained low in concentration. As a result the amount of nitrogen and phosphorus that reached the NW Black Sea was much higher and the amount of silicate was much lower than before the 1970s (Humborg et al., 1997). Consequently, the silicon to nitrogen ratio became much lower. After the 1970s coastal waters experienced massive blooms of microalgae that consisted of dia- toms, dinoflagellates, euglenophytes, and prymnesiophytes, while formerly only diatoms and dinoflagellates were present (Humborg et al., 1997). This had further effects on food webs and physicochemical and biogeo- chemical properties of water and sediment (Mee et al., 2005). Similarly, the combination of river damming and land-use change in the Mississippi River basin led to a decrease of the Si:N ratio that also altered the phyto- plankton community structure in the Gulf of Mexico, mostly in the sense that diatoms became less silicified (Rabalais et al., 2000). Coastal eutrophication can also manifest itself in massive macroalgal blooms that may have a remote origin. The world’s largest macroalgal 66 Microbial Communities in Coastal Sediments blooms of Ulva prolifera were observed off the Chinese city of Qingdao between 2008 and 2012 (Fig. 2.12; Liu et al., 2013). Approximately 400 km2 of the Qingdao coast was covered by about 1 million tonnes of algae biomass. Most probably, the problem originated from aquaculture facilities that grow Porphyra seaweeds along the Jiangsu coast, hundreds of kilometers south of Qingdao. U. prolifera is a fouling green alga commonly found on Porphyra mariculture rafts, which is frequently being removed from them. The large-scale cleaning of rafts released about 5000 tonnes of wet weight of U. prolifera into coastal waters. From there, small patches were transported to the North with a coastal current and grew to massive carpets of algae while traveling through nutrient-rich and warm waters supporting their growth (Liu et al., 2013). Nixon (2009) mentioned the importance of the macroscopic view on eutrophication phenomena, that is, to see them in larger scales, like the abovementioned example has demonstrated. However, the microscopic view is also required in order to detect small-scale changes, the cumulative effects of which may have deleterious effects on timescales of decades to centuries. Small low-lying islands in tropical regions come into focus mostly because of their vulnerability to sea-level rise. However, their “effluent aura” may also contribute to coastal eutrophication. A

Figure 2.12 2008 Summer green tide in the Yellow Sea, China: (A) geographic location of Qingdao and (B) photograph on June 27, 2008 showing green tide at Qingdao coast. Taken from Liu, D., Keesing, J.K., He, P., Wang, Z., Shi, Y., Wang, Y., 2013. The world’s largest macroalgal bloom in the Yellow Sea, China: formation and implications. Estuar. Coast. Shelf Sci. 129, 210. Sources, types, and effects of nutrients (N and P) in coastal sediments 67 comparative study of two such islands, the densely populated Pulau 2 Barrang Lompo (4000 inhabitants on 0.2 km2 5 20,000 inh km 2) and the uninhabited Pulau Kodinggareng Keke in the Spermonde Archipelago, Sulawesi, Indonesia, found statistically significant elevated phosphate and Chl a levels in surface waters up to 300 m offshore of the inhabited island that lacks proper sewage and wastewater treatment facili- ties. However, in both cases nutrient and chlorophyll concentrations were below commonly applied threshold values of ,19 μM TN, ,0.4 μM 2 TP, and ,1 μgL 1 Chl a (Smith et al., 1999; Kegler et al., 2018) and as such qualified as “oligotrophic.” Obviously, the effluents from this densely populated point source are rapidly diluted in the large sea area around the 0.2 km2 island. Hence, the microscopic view indicates that human activi- ties have already modified coastal waters, but commonly applied macro- scopic evaluation schemes do not indicate the need for action. Anthropogenic eutrophication is also impairing the well-being of sea- grass meadows and coral reefs, two other ecologically and economically important coastal ecosystems (e.g., Hughes et al., 2003; Orth et al., 2006). Eutrophication is the most significant and most common cause of seagrass decline that has been documented both in temperate and tropical coastal ecosystems (Duarte, 2000; Waycott et al., 2009). Seagrasses are efficient nutrient recyclers as they require only little nutrients for their growth and are adapted to oligotrophic conditions. In seagrass meadows, sewage and aquaculture induce eutrophication and a strong inverse relationship has been shown between eelgrass habitat and nitrogen concentration in the overlying water (Olesen, 1996). As eutrophication proceeds, seagrasses can be replaced by algae. Consequently, in the sediment, long-term retention of recalcitrant, dissolved and particulate matter will decline, which will be replaced by less refractory material. Due to the movement of ephemeral algae along with water, mass transport of plant-bound nutrients will increase. For N, denitrification will become an unimportant sink, because of increased primary production in the water column that will outcom- pete denitrifiers for available N. As a result, in the later stages of eutrophi- cation, nitrate reduction will shift to DNRA (McGlathery et al., 2007). On the Chinese island of Hainan the massive aquaculture effluent release directly into coastal back-reef areas strongly impaired seagrasses and coral reefs. The use of stable nitrogen isotope measurements allowed to trace the anthropogenic nitrogen from its source as ammonium in aqua- culture pond waters until its sink in seagrass tissue. The isotopically heavy 15 ammonium (δ N-NH4 15m18m) entering coastal waters was taken up 68 Microbial Communities in Coastal Sediments by microalgae and seagrasses leading also to elevated δ15N values of 8m12m and 7m9m, respectively, the latter of which are among the highest measured worldwide (Herbeck and Unger, 2013; Herbeck et al., 2014). Massive growth of microalgae in the water and of epiphytic algae on the seagrasses strongly impaired the well-being of the latter. Gradients of seagrass performance were observed with a decreasing abundance and diversity of seagrasses with increasing exposure to aquaculture pond efflu- ents (Herbeck et al., 2014). Adjacent coral reefs in the area were generally in a very bad condition, with coral live cover mostly being in the range of 5%10% in five coral reefs along the NW Hainan coast. However, eutrophication was only one among the several factors impairing coral health, with overfishing being a major factor (Roder et al., 2013).

2.4.2 Hypoxia and anoxia in water and sediment The most common and obvious effect of eutrophication is depletion of dissolved oxygen leading to hypoxia and anoxia in the water column and in the sediment. Hypoxic and anoxic waters differ qualitatively as well as in the quantity of oxygen they contain. While anoxic waters are free of oxygen, there is no universally applicable and accepted definition of “hyp- oxia.” The term usually refers to water masses that are undersaturated with dissolved oxygen and is mostly used in context with restrictions of the performance of aquatic organisms. A broadly used threshold value is 2 2mgL 1, a concentration below which physiological functions of many organisms are impaired and which refers to the threshold value for fisher- ies collapse (Renaud, 1986). However, it has been demonstrated that hyp- oxia tolerance and threshold values vary largely. The behavioral, physiological, and reproductive responses differ by species, stage of life, and the length of exposure to low oxygen levels (Ekau et al., 2010; Rabalais et al., 2010). Experimentally derived oxygen thresholds for lethal and sublethal responses to hypoxia vary by an order of magnitude among benthic organisms. For example, for some fish and crustacean species 2 lethal oxygen concentrations were in the range of 89mgL 1, while 2 they were ,1mgL 1 for some gastropod species (Vaquer-Sunyer and Duarte, 2008). Hypoxia in bottom waters of coastal ecosystems is typically caused when large amounts of algae die, sink to the bottom, and are decomposed by bacteria that use up the available oxygen. It frequently happens, for example, in the Gulf of Mexico that receives the nutrient-laden waters of Sources, types, and effects of nutrients (N and P) in coastal sediments 69 the Mississippi River that stimulate excessive algal growth. The microbial decomposition of this large biomass during summer usually leads to 20,000 km2 large area of bottom water with dissolved oxygen concentra- 2 tions of ,2mgL 1, which is often referred to as “dead zone” (Rabalais et al., 2010). Although hypoxia is typically seasonal, some systems such as the Baltic and Caspian seas experience year-round hypoxia due to the severity of eutrophication (Karlson et al., 2002). This process differs between coastal ecosystems, depending on geomorphological features of the estuaries and activities in the hinterland that influence the amount and composition of available nutrients and the formation of benthic hypoxia. For instance, damming of the Danube River and land-use change in Romania led to a large increase of nitrate and phosphate inputs and a large decrease in silicate inputs into the NW Black Sea. This caused a large increase in algae blooms and a shift in community composition toward more nonbiomineralizing species in the 1970s and 1980s. While the biomineralizing diatoms remained the dominant phytoplankton group, the much lower Si:N ratio favored the growth of other, partly nonbiomineralizing species. This has implications for food webs, biogeo- chemistry, and oxygen levels. While biomineralizing plankton rapidly sink to the bottom and are mainly decomposed there, nonbiomineralizing plankton are recycled to a larger extent in the upper water column. However, in the case of the NW Black Sea shelf the very massive blooms led to large zones of bottom-water hypoxia and also to a loss of benthic biodiversity and changes in the benthic community composition, for example, from bivalves to polychaetes (Friedrich et al., 2014). Benthic hypoxia can cause adverse impacts farther up the food web such as changes in benthic biodiversity and (Boesch, 2002). Hypoxia and anoxia can change the makeup of a community by killing off more sensitive or less mobile organisms, reducing suitable habitat for others, and changing interactions between predators and their prey. For instance, recurring periods of low oxygen tend to shift the dominance of benthic com- munities in the coastal sediment away from large, long-lived species such as clams to smaller, opportunistic, and short-lived species such as polychaete wormsthatcancolonizeandcompletetheirlifecyclesquicklybetweenthe periods of hypoxia (Pearson and Rosenberg, 1978). Hypoxic and anoxic conditions in the sediment can also largely affect benthic flora and fauna (e.g., Raven and Scrimgeour, 1997). It may lead to massive die-offs as, for example, reported from the Black Sea (Fig. 2.13). Moreover, oxygen depletion in the sediment can also affect 70 Microbial Communities in Coastal Sediments

Figure 2.13 Photograph of a sediment sample taken from the NW Black Sea shelf off the Danube River mouth during a research cruise in 1999. The sediment is anoxic, there was a smell of hydrogen sulfide, no live animals could be seen upon visual inspection. Photograph taken by T.C. Jennerjahn. the well-being of seagrasses. In such a situation decomposition is often dominated by anaerobic pathways like, for example, sulfate reduction. It can lead to sulfide accumulation in the sediments and high concentrations of toxic hydrogen sulfide in porewater (Holmer and Kristensen, 1992; Frederiksen et al., 2007; Holmer and Frederiksen, 2007). In coastal waters off the Chinese island of Hainan, sulfur isotope measurements indicated the uptake of toxic hydrogen sulfide by seagrasses. There, coastal eutro- phication related to aquaculture effluents also stimulated microbial decom- position of the enhanced sedimentary organic matter. As a result of oxygen depletion the further decomposition of organic matter required the consumption of other oxidants, for example, sulfate. While roots and 34 rhizomes are usually capable of oxidizing the H2S, the low δ Sof,21m in the tissue of seagrass leaves off Hainan indicated the uptake of Sources, types, and effects of nutrients (N and P) in coastal sediments 71 hydrogen sulfide from the sediment and transport into the leaves (Herbeck et al., 2014).

2.4.3 Eutrophication-induced changes in sediment microbial communities Eutrophication causes increased sedimentation of organic matter to the coastal sediments and in systems of recurring hypoxia, more organic mat- ter is available for remineralization by microbial communities (Boesch, 2002). The effect of nutrient enrichmentinduced eutrophication results in increased supply of organic matter that stimulates microbial populations and processes in the sediment, thus elevating the microbial loop to a more prominent trophic pathway. The responses of eutrophication may be acute or chronic, subtle or profound, depending on several factors. For example, light may be a limiting factor in increasing primary production of phytoplankton with increased availability of N and P. Increased phyto- plankton biomass, however, in turn reduces water transparency that results in increased delivery of organic matter to the sediments. In temperate ecosystems, nutrient enrichment does not stimulate perennial grass populations as these plants obtain their nutrient require- ments from stored nitrogen pools and also from sediment (Pedersen and Borum, 1996). Hence, nutrient enrichment causes blooms of phytoplank- ton and macroalgae. In tropical systems also fast-growing “nuisance” macroalgae and phytoplankton with high nutrient uptake potential rapidly replace seagrasses. The presence of macrophytes such as seagrasses buffers against external nutrient loading, acting as “coastal filters.” They also transfer oxygen and thus actively oxygenate the bottom sediments, retain- ing P as ferrophosphates and creating suitable conditions for nitrification and denitrification. Moreover, the presence of roots of macrophytes cre- ates a rhizosphere with a special microhabitat for bacterial activity and oxi- dation of iron and sulfur and also fixing phosphorus within the sediment. Thus obligate anaerobic conditions are prevented by active pumping of oxygen and the strength of this depends on the amount of photosynthesis (Borum et al., 2005; Duarte, 2000). On the other hand, eutrophication- induced phytoplankton and macroalgal blooms increase the organic matter in the sediment of seagrass meadows and reduce oxygen translocation and release to the rhizosphere. Thus exceeding the tipping point of sediment oxygenation is the primary factor in the shift to a phytoplankton- dominated sate. 72 Microbial Communities in Coastal Sediments

Denitrification is a major sink for nitrogen in coastal ecosystems and it tends to drive systems toward nitrogen limitation. The overall magnitude of denitrification is greater in estuaries than in freshwater ecosystems due to greater nitrogen fluxes through estuaries. However, systems with longer water residence time experience more nitrogen loss through denitrifica- tion (Howarth et al., 1996; Nixon et al., 1996a). A sediment process that counteracts the influence of denitrification on nutrient limitation is phos- phorus adsorption. For example, in the Narragansett Bay sediment, little or no phosphorus is adsorbed by the sediments and all of the phosphate produced during decomposition in the sediments is released to the water column (Nixon, 1997). Hence, this bay is nitrogen limited since the phos- phate in combination with nitrogen is lost through denitrification. Eutrophication may lead to less denitrification since the coupled processes of nitrification and denitrification are disrupted in anoxic waters. Similar studies in the Chesapeake Bay sediments show an intermediate behavior with some of the phosphorus, released during sediment decom- position, being absorbed and some released back to the water column (Boynton and Kemp, 1985). The ability of both tropical and temperate coastal marine sediments to adsorb and store phosphorus decreases as they become more eutrophic. The difference is that, for temperate systems, the phosphate sorption is less due to lower amounts of oxidized iron and more iron sulfides in the sediments. However, in tropical carbonate sys- tems, the rate of sorption of phosphate decreases as the phosphorus con- tent of the sediment increases. Increased phosphorus availability in eutrophic systems intensifies nitrogen limitation and stimulates growth of algae and other organisms with high phosphorus requirements. Thus eutrophication and pelagic nutrient enrichment affect the coastal marine sediments by increased sedimentation of organic matter and poor benthic oxygenation. This causes a shift in the dominance of heterotrophic micro- bial food webs over the “classic” planktonic food chain (Korpinen and Bonsdorff, 2015). Increased anoxic conditions also favor the predomi- nance of obligate anaerobes like, for example, sulfate-reducing bacteria and methanogenic archaea in the sediments. This occurs as a result of high oxygen consumption rates in the water column and low oxygen transport by molecular diffusion, thus gradually limiting the depths of oxygen penetration in the sediment. The oxygen penetration depths may be a few millimeters in fine sediments (muddy) and a few centimeters in permeable coarse sediments (sandy). Ultimately, a cascade of electron acceptors used by various groups of microorganisms that occupy different Sources, types, and effects of nutrients (N and P) in coastal sediments 73

Figure 2.14 Nutrient enrichment and eutrophication-induced changes in sedimen- tary microbial structure. ecological guilds is initiated by oxygen availability or limitation. As oxy- gen gets depleted and the system moves from oxic to hypoxic conditions, aerobic microorganisms are replaced by facultative anaerobes such as deni- trifying, manganese-reducing, and iron-reducing bacteria that utilize nitrate, manganese, and iron hydroxides as the terminal electron acceptors. Further from hypoxic to anoxic conditions, sulfate reducers and methano- gens predominate under obligate or strict anaerobic conditions (Fig. 2.14). This cascadal sequence is governed by thermodynamic yield, reaction kinetics, and the physiology of the microorganisms involved.

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3.1 Introduction Coastal marine sediments are complex systems influenced by the interac- tions of various geological, hydrological, physical, chemical, and biological factors (Zhang et al., 1999). The striking feature of microbes in coastal sediments is their extreme diversity that is due to the fact that the specia- tion rate is faster than the extinction rate (Dykhuizen, 1998) and generally high speciation rates are observed. Hence, it is important to investigate how microbial communities vary over spatial scales, temporal scales as well as the influence of other external factors on microbial diversity and abundance (Prosser et al., 2007). Although, in a broader perspective, geo- graphical distance and environmental variables are the two main factors that drive microbial distribution patterns in sediments (Martiny et al., 2011), microbial community composition and activity are greatly influ- enced by an assortment of biotic and abiotic factors. Those include physi- cochemical characteristics of the sediment as well as overlying water, availability and type of electron donor and acceptor, competition for space and resources, etc. Furthermore, differences in rates and diversity of microbial activity have also been linked to changes in community compo- sition (Castro et al., 2010; Fierer et al., 2012). In coastal environments, despite a constant influx of input from the overlying water column (Morris et al., 2002), there exists a clear differ- ence between planktonic and benthic communities, with significant changes in benthic community composition (Zinger et al., 2011; Walsh et al., 2016). This delineation in benthic community structure from other marine compartment suggests that unique and decisive processes are responsible for driving the community composition of sediment environ- ments. Among several processes, the factors regulating benthic diversity are particularly the relationship between diversity and resources availabil- ity/productivity (diversityenergy relationships) that has rarely been

Microbial Communities in Coastal Sediments © 2021 Elsevier Inc. DOI: https://doi.org/10.1016/B978-0-12-815165-5.00003-0 All rights reserved. 79 80 Microbial Communities in Coastal Sediments investigated (Langenheder and Prosser, 2008). In addition, links between environmental parameters and community composition have been pro- posed (Inagaki et al., 2006; Jorgensen et al., 2012); however, these appear to be overshadowed by global distribution of dominant populations that occurs independently of environmental conditions (Orcutt et al., 2011; Parkes et al., 2014). Although microbes play a key role in ecosystem func- tioning, the mechanisms controlling biodiversity and function in microbial communities are still to be elucidated. Vellend (2010) proposed a conceptual synthesis that centers on four major assembly mechanisms: (1) dispersal, that is, the movement of cells across space; (2) speciation (or diversification), that is, the generation of genetic variation; (3) drift, that is, random changes in community compo- sition across time; and (4) selection, that is, the change in community structure due to fitness differences among community members. In addi- tion, biological interactions also control the local diversity in coastal sedi- ments. However, none of these processes act in isolation and the relative importance of each may vary among different systems. Nevertheless, this framework has provided a unified lens through which microbial commu- nity assembly patterns can be investigated (Vellend, 2010; Nemergut et al., 2013; Stegen et al., 2013; Fig. 3.1). Several studies have found that the major environmental factors that affect bacterial community composition and structure in coastal sediments are often salinity (Bouvier and Giorgio, 2002), nutrients (Biddanda et al., 2001), pH (Zheng et al., 2014), substrate types (Li et al., 2014), and sedi- ment grain size (Wang et al., 2017). This chapter discusses the various biotic and abiotic factors that influence microbial structure and function in coastal sediments. The overall factors are given in Fig. 3.2.

3.2 Spatial and temporal heterogeneity Microbial cells rarely show uniform distribution in the environment, because microenvironments that form within niches determine the resource availability to microbial cells. Coastal sediments are chemically complex with steep gradients of redox potential, pH, and substrate avail- ability that contribute to the formation of large number of microhabitats. The structural heterogeneity of sediments allows resource partitioning, thus creating new niches and enhancing diversification into distinct eco- logical species (Torsvik et al., 2002). Spatial changes in microbial commu- nity structure and abundance also depend on dispersal. Dispersal is the Environmental variables and factors regulating microbial structure and functions 81

Figure 3.1 Microbial community assembly mechanisms in sediments (Petro et al., 2017).

Figure 3.2 Factors influencing sediment microbial community structure. movement of organisms across space, which is of two types passive or active (Nemergut et al., 2013). Active dispersal is maximum in the biotur- bation zone (burrowing activity of animals), due to homogeneity of the environment and increased availability of freshly deposited organic matter 82 Microbial Communities in Coastal Sediments

(Fenchel, 2008). On the other hand, passive dispersal is restricted to bio- turbation zone, where burrowing activity by benthic macrofauna con- stantly redistributes microbial cells along with particulate matter (Meysman et al., 2005). The structure of sediment matrix also limits this process as bacterial cells have a tendency to adhere to sediment particles (Lever et al., 2015). Dispersal due to motility serves as a means to avoid burial in locations with low sedimentation rate. Considering the fact that microbial cells are rarely uniformly distributed in the environment, spatial scales are important. This is because microenvironments form within niches that determine the availability of resources to microbial cells. Temporal scales are important for microbial communities because with short-generation times, microbes show rapid growth and maximum growth is attained where conditions are favorable. Changes in environ- mental conditions as well as resource availability over time are the major factors that create opportunities for new species to become established. Eventually, this will lead to varied patterns of microbial diversity over dif- ferent temporal scales and evolution in microorganisms can occur rapidly with the convergence of ecological and evolutionary timescales (Prosser et al., 2007). Degradation of organic matter over time also changes the composition of the resource caused by the , which in turn changes the succession of microbial community over time and space. Consequently, microorganisms with a potentially high growth rate (r-stra- tegists) will become numerically dominant and reduce the evenness of the species distribution (Torsvik et al., 2002). Succession of community also happens over time and where organic matter is being degraded, changes in the composition of the resource caused by the microbial metabolism change the community over time and space. Hence, future studies would be directed to understand the dynamics of the microbial populations at a higher spatial and temporal resolution.

3.3 Geological factors 3.3.1 Sediment granulometry Sediment granulometry (sand, silt, and clay content) is an important factor in shaping bacterial community composition in the sediments, with sand content showing an influence on community richness and diversity (Zheng et al., 2014). Sandy or coarse sediments cover large areas of coastal environment and have a large role in global biogeochemistry. The high Environmental variables and factors regulating microbial structure and functions 83 permeability of coarse sediments allows for rapid exchange of pore water with the overlying water column. Thus these sediments rapidly recycle organic matter and play a large role in biogeochemical cycles. Sandy sedi- ments also enhance the transport of microbial substrates into the sediments and also metabolic waste products out of the sediments. Sandy sediments maintain a low microbial abundance due to enhanced hydrodynamic activity due to transport and exchange (Rusch et al., 2003). Due to differ- ent physicochemical composition, microbial community structure is dif- ferent for fine- and coarse-grain sediments. Contrastingly, molecular diffusion limits aerobic as well as anoxic microbial metabolism to a thin fine layer in fine-grained sediments. Liu et al. (2015) found that the sedi- ment median size content could regulate the microbial community in the four marginal seas of northern China.

3.3.2 Sediment depth 3.3.2.1 Shift in substrate availability with depth In coastal sediments, there is an increasing substrate limitation of the active bacteria with depth, because dissolved organic carbon (DOC), including dissolved carbohydrates, becomes increasingly unavailable to the microor- ganisms with increasing depth. This results in vertical depth gradients of organic compounds that serve as energy sources for microorganisms. Studies conducted to examine the effect of energy gradients on the com- position of the microbial community revealed a significant shift in the microbial community structure with depth. An ecological classification of bacterial phyla into copiotrophic and oligotrophic categories has been observed in the sediments relative to substrate availability (Fierer et al., 2007). It is evident that the abundance of active bacteria is directly related to the availability of electron acceptors and energy sources within the sediment and decreases with depth. In contrast, the phylogenetic richness and diver- sity may increase with depth, despite decrease in availability of electron acceptors and energy sources with sediment depth. This is due to the presence of microorganisms that are adapted to thrive in abundant carbon sources that occur in the surface part of sediment. In contrast, high bacterial diversity and low abundance are observed in the bottom sediment where carbon sources become increasingly unavail- able to microorganisms. Microorganisms adapted to low amounts of read- ily degradable carbon sources are referred to as oligotrophs and they have an allochthonous mode of feeding (Koch, 2001). They exhibit K-strategist life histories with a low growth rate (Fierer et al., 2007). 84 Microbial Communities in Coastal Sediments

However, the species richness of microorganisms within the environment is controlled by a multitude of physical and biological factors. For exam- ple, a study of the equatorial Pacific Ocean and North Pacific Gyre showed a decrease of the total number of species with depth, which was evident from the Chao 1 index that provides an estimate of species rich- ness, taking into account the number of rare species present within a sam- ple (Hughes et al., 2001; Walsh et al., 2016). Hence, it can be inferred that not all microorganisms in the top sediment are copiotrophs and simi- larly not all microorganisms in deeper sediments are oligotrophs. Globally, such reduction in diversity and richness indicates selection suggesting disappearance of certain population as they are buried deeper in the sediment. The imposed energy limitations remove specific popula- tion as well as reducing the size of total microbial community that reflects the energy-driven drop in microbial count with depth (Kallmeyer et al., 2012; Jørgensen and Marshall, 2016). The possible adaptation by microor- ganisms to overcome energy depletion in the deeper sediments is by increasing ATP synthesis efficiency. Those microorganisms that lack this adaptation undergo sporulation in order to be prevalent throughout the sediment depth, thus allowing them to switch to dormancy when there is energy depletion in the sediments (Petro et al., 2017).

3.3.2.2 Decrease in lability of organic matter with depth The quality of organic matter throughout the sediment depth is an impor- tant determining factor for bacterial growth and function, which ulti- mately affects the bacterial composition and phylogenetic richness throughout the sediment depth. Lability of DOC is more in the surface sediment and becomes increasingly resistant to degradation with depth, probably due to the formation of complex polysaccharides and polysul- fides. The amount of hydrolyzable DOC decreases with increasing decomposition and diagenetic alteration throughout the sediment depth (Cowie and Hedges, 1994). DOC that resists degradation accumulates with depth in sediment (Burdige and Gardner, 1998; Hedges, 1988). Microorganisms adapt to exploit the various pools of organic carbon throughout the sediment depth indicating a vertical niche separation. Presence of bacterial phylotypes in the surface layer and absence in the bottom layers reflect the specialization of these phylotypes to specific eco- logical niches within the sediment. Thus the partitioning of resources within the sediment creates specific niches, enhancing microbial specializa- tion and division into distinct ecological guilds (Torsvik et al., 2002). The Environmental variables and factors regulating microbial structure and functions 85 higher amount of refractory organic carbon in the deeper sediment makes it an environment with limited carbon resources and the resultant compe- tition for the limited, accessible DOC within the deeper sediment may lead to niche specialization and diversification. This results in higher phy- logenetic diversity in the deeper sediment. Moreover, microbial response to the available DOC in the sediment depends on the physiological capa- bilities of the microorganisms that ultimately determine which microor- ganisms can coexist successfully within the environment (Madrid et al., 2001). Members of the Bacteroidetes and Gammaproteobacteria possess enzymes for polysaccharide hydrolysis and fermentation (Alderkamp et al., 2007; Kirchman, 2002) and the distribution of these bacteria throughout the depth of the sediment can give an initial indication of a possible link between hydrolysis rates and the abundance of potential hydrolytic and fermenting bacteria. The higher metabolic rates in the surface sediment are matched by a higher abundance of Bacteroidetes, Gammaproteobacteria, and sulfate-reducing bacteria (SRB), compared to the deeper sediment. The high diversity of SRB throughout the sediment depth is a result of niche partitioning, which is the way in which the dif- ferent species distribute the available resources in the environment. According to the “niche overlap hypothesis” (Pianka, 1974), highly diverse communities arise in environments, which are stable over long periods of time as a result of competition-maintained niche diversification.

3.4 Hydrological factors Spatial differences in biogeochemical rates and bacterial abundance are attributed to differences in hydrographic features, such as current strength and direction, which cause enhanced sediment reworking in shallower regions. This in turn has an effect on bulk sediment composition, affecting the amount of particulate organic carbon and nitrogen within the sedi- ment. Highest microbial activity and metabolism take place in regions where the sediment is well mixed due to hydrological influence.

3.5 Physicochemical factors Among the various physicochemical factors that influence microbial distri- bution and availability, the predominant factors are pH, salinity, redox potential, organic carbon, and nutrients. Various studies reveal that none 86 Microbial Communities in Coastal Sediments of the individual physiochemical properties significantly affected activity or community composition. However, the complex interactions among the various sediment properties drive the significant differences in micro- bial activity and community composition. The chemistry of overlying water can also directly influence the benthic microbial communities, par- ticularly in the surface layer of the sediments.

3.5.1 pH Although benthic microbial communities change with changes in pH in coastal sediments (Tait et al., 2013), pH may not be a significant factor that influences bacterial abundance, diversity, and community structure in the coastal sediments (Zheng et al., 2014). However, pH was identified as a significant factor that determined community distribution of sediment bacteria in freshwater systems such as rivers (Liu et al., 2015). pH is a uni- versal indicator of bacterial community structure for alkaline lake sedi- ments (Xiong et al., 2012) with dominance of Alphaproteobacteria in higher pH and Acidobacteria in low pH (Chu et al., 2010).

3.5.2 Salinity Salinity is a major driver of microbial community composition and func- tion in coastal sediments. Overall shift with changing salinity is observed for microbial structure as well as functions including amino acid and car- bohydrate metabolism in the sediments (Kimbrel et al., 2018). The influ- ence of salinity on sediment microbial communities is more pronounced in coastal ecosystems with a salinity gradient like estuaries. Some of the taxa such as Bacteroidetes and Proteobacteria dominated higher salinity communities, which are attributed to their ability to degrade complex organic matter (Blümel et al., 2007; Campbell and Kirchman, 2013; Dupont et al., 2014). Gammaproteobacteria can dominate brackish habi- tats and are favored by high due to their opportunistic life strate- gies (Pavloudi et al., 2016, 2017; Zhang et al., 2014). Salinity has an important role in controlling the occurrence and diversity of archaea in sediments. For example, among the different types of methanogenic archaea (MA) based on substrate utilization, low-salinity sediments were dominated by acetotrophic and hydrogenotrophic methanogens. Whereas, methylotrophic methanogens were more prevalent in the highly saline sediments (Webster et al., 2015). Microbial activities are controlled by Environmental variables and factors regulating microbial structure and functions 87 salinity due to energetic constraints and rare taxa are more sensitive to salinity than abundant taxa (Yang et al., 2016).

3.5.3 Pore water chemistry/presence of nutrients or chemicals 32 22 The bioavailable forms of phosphorus (PO4 and HPO4 ), nitrogen 1 2 2 2 (NH4 ,NO2 , and NO3 ), and carbon (CO2, HCO3 , and carbohydrates) are commonly used by microorganisms and their availability influences and shapes bacterial community composition in the sediments (Jansson et al., 2006; Hou et al., 2015; Gao et al., 2016; Soares et al., 2017). Increased ammonia results from bacterial decomposition, mineralization, and ammonification of dissolved organic nitrogen derived from plant detritus in reduced sediments (Kristensen et al., 1988). Increased concen- trations of ammonia in pore water dominate the benthic nitrogen in estu- arine and mangrove sediments. Low pore water concentrations of ammonia indicate that the turnover of inorganic nitrogen is limited and result in tight coupling between microbial assimilation and decomposition (Alongi, 1996; Kristensen, 1988). Similarly, more dissolved phosphate in aerobic sediments is explained by the increased phosphate input from degrading leaf litter mangroves. Higher phosphate and nitrogen content in sediment may favor the nitrogen-fixing bacteria (Sjoling et al., 2005). Total phosphorus in sediment was found to be significantly correlated with the bacterial community (Biddanda et al., 2001).

3.5.4 Redox potential Redox potential is also a key environmental factor shaping microbial community structure and function (DeAngelis et al., 2010). Reductionoxidation (redox) state is a chemical characteristic that influ- ences early diagenesis and the burial of organic carbon in coastal sediments (Reimers et al., 2013). Sediment redox state is codependent with other factors such as deposition rate, OM flux, benthic faunal activities, and bot- tom water oxygen concentration. When sediment is exposed to oxygen and attains positive redox conditions, oxidation of sedimentary organic matter including more refractory OM pools occurs over a period ranging from annual to millennial timescales (Burdige, 2007). Oscillating redox conditions, which are geochemically complex occurrences, are also widely observed in coastal sediments due to interactions by macrofauna and meiofauna like bioturbation and bioirrigation (Aller et al., 2001). 88 Microbial Communities in Coastal Sediments

Moreover, physical mixing events can also cause bed fluidization that results in shift in redox state of sediments (Aller et al., 2008).

3.5.5 Changes in availability of electron acceptors During decomposition of organic matter, complex organic compounds are converted to simpler compounds and further mineralized by microor- ganisms in the sediment. During this process, microorganisms derive energy by transferring electrons from an external electron source or donor to an external electron sink or terminal electron acceptor (Megonigal et al., 2003). Bacterial communities in sediments are metabolically diverse and can utilize various electron acceptors, including oxygen, nitrate ð 2Þ NO3 , manganic manganese [Mn(IV)], ferric iron [Fe(III)], sulfate ð 22Þ SO4 , and carbon dioxide (CO2)(Lin et al., 2006; Nakagawa et al., 2005). The fundamental difference between aerobic and anaerobic metab- olism is energy yield (Gibbs free energy, ΔG°) (Table 3.1). Oxidation of glucose yields more energy under aerobic conditions, whereas under obligate anaerobic conditions such as methanogenic conditions, the energy yield is less. However, the high-energy yield permits a single aero- bic microorganism to completely oxidize complex organic compounds to CO2. Contrastingly, under anaerobic conditions, mineralization of organic carbon to CO2 is a multistep process that involves a consortium of micro- organisms (Fig. 3.3). Those that live in deep anoxic waters as obligate or facultative che- moorganotrophs survive by fermenting organic matter (Lin et al., 2006). The availability of TEAs highly influences the terminal step of anaerobic decomposition (Capone and Kiene, 1988; Megonigal et al., 2004) by act- ing through interspecific competition for electron donors (acetate and H2). The ability of microorganisms to use multiple electron acceptors is an advantage for them to remain active in an environment where the sup- ply of specific electron acceptors is variable (Megonigal et al., 2003; Table 3.1). Although the importance of various TEA pathways has been much reported, only little work has been done on the mechanistic control of overall anaerobic mineralization of organic matter by both electron donor ð 22Þ and acceptor supply. Increased sulfate SO4 loading from sea level rise 2 (Weston et al., 2006) and increased NO3 loading from cultural eutrophi- cation (Vitousek et al., 1997) or acid deposition (Wieder et al., 1990) can alter both electron donor and acceptor availability in coastal ecosystems. Table 3.1 Redox reactions and free energy yield during various electron accepting processes in the sediments (Jørgensen, 2000; Megonigal et al., 2004). Sl. No. Electron accepting process Free energy yield Eh (V) 2 ΔG° (kJ mol 1) 1 Aerobic respiration 2479 0.812 CH2O 1 O2-CO2 1 H2O 2 Nitrate reduction 2453 0.747 1 2- 1 2 1 1 CH2O 4NO3 2N2 4HCO3 CO2 H2O 3 Manganese (IV) reduction 2349 0.526 1 1 1 - 1 1 2 CH2O 3CO2 H2O 2MnO2 2Mn2 4HCO3 4 Iron (III) reduction 2114 20.047 1 1 ðÞ- 21 1 2 1 CH2O 7CO2 4Fe OH 3 2Fe 8HCO3 3H2O 5 Sulfate reduction 277 20.221 1 22- 1 2 CH2O SO4 H2S 2HCO3 6 Methanogenesis 228 20.244 2 1 CH3COO 1 H -CH4 1 CO2 90 Microbial Communities in Coastal Sediments

Figure 3.3 Vertical organization of electron acceptor availability.

Moreover, sea level riseinduced sulfate loading (the concentration of sulfate in seawater is B28 mM) can lead to a shift of electron flow from methanogens to the sulfate reducers that are competitively dominant than methanogens (Neubauer et al., 2005; Weston et al., 2006). Methanogens and SRB compete for hydrogen and acetate, SRB have higher affinity for H2 than methanogens and hence, outcompete them (Fig. 3.4). However, those SRB that utilize acetate as substrate have a slightly more advantage thermodynamically and take longer time to outcompete methanogens (Stams et al., 2005).

3.5.5.1 Aerobic respiration Approximately 50% of the organic matter oxidation in marine sediments is considered to take place under aerobic conditions (Canfield, 1993). Dissolved oxygen content regulates microbial community and organic matter that reaches the sediments is aerobically respired until it exceeds the amount of oxygen that can be delivered to the site of diffusion. High organic content in the sediment favors processes of decomposition that deplete the sediment of oxygen, resulting in more anoxic conditions (Sjoling et al., 2005). Consequently, changes in the sediment oxygen levels can affect the diagenetic pathways, which can turn more anaerobic at the expense of aerobic pathways (Middelburg and Levin, 2009). Oxygen penetration in the sediments varies from 1 mm in the fine Environmental variables and factors regulating microbial structure and functions 91

Figure 3.4 Sulfate reduction and methanogenesis in the presence (red lines) and absence of sulfate. sediments to few centimeters in the sandy sediments that are permeable. Following oxygen depletion, anaerobic respiration is initiated by using a cascade of alternative electron acceptors such as nitrate, manganese and iron (hydr)oxides, and sulfates.

3.5.5.2 Nitrate reduction Coastal sediments are important sites for processing nitrate from water col- umn and nitrate in sediments is quickly turned over and becomes depleted when it is no longer recharged by transport of nitrate from the water col- umn (Robinson et al., 1998). Nitrate is often limiting at the marine sites of estuaries although hypernutrified estuaries show a high turnover of nitrate where it is constantly resupplied. For example, the Colne estuary shows strong gradients of nitrate and ammonium from the estuary head to the mouth where 20%25% of the total N load entering the estuary is removed by denitrification (Dong et al., 2000). While proceeding down- ward across the oxygen depletion zone, nitrate concentration increases 92 Microbial Communities in Coastal Sediments due to nitrification process. Denitrification is the most common form of anaerobic respiration based on nitrogen. Respiratory denitrification is more energetically favorable than Fe(III) reduction, sulfate reduction, or methanogenesis (Megonigal et al., 2004). Nitrate reduction is mediated by a diverse polyphyletic group of bacte- ria (Zumft, 1997) due to the fact that each bacterial species may partici- pate in only one step of the denitrification process. In addition, for denitrifying bacteria, the dominant environmental factor affecting the community composition and structure is nitrogen content (Hou et al., 2014; Zheng et al., 2014). Geographical location can affect the distribu- tion and diversity of denitrifiers in sediments (Gao et al., 2016). Denitrification is primarily dependent upon anoxic conditions, availability of nitrate, and the presence of an energy source. Evidence suggests that only 27%57% of the reduced nitrate is accounted for by denitrification; the remainder is reduced to ammonium by dissimilatory pathways (Koike and Hattori, 1978). This fraction may, however, vary according to sedi- ment type and redox regime. Increased denitrification in the oxic surface sediment may imply the presence of anaerobic microniches that favor high activity of chemoheter- otrophs in the oxic zone (Kristensen, 1985). In surface sediments, how- ever, only 6%35% of the nitrate produced generally is consumed by denitrification (Billen, 1978; Nishio et al., 1982). Denitrification is limited 2 by availability of NO3 and hence may be coupled to nitrification, which 2 is production of NO3 . Therefore it is conceived that a major source of 2 NO3 in the sediments is by nitrification than that enters into the sedi- ment by diffusion from the overlying water (Salahudeen et al., 2018).

3.5.5.3 Mn and Fe reduction Microorganisms that reduce extracellular Fe(III) or Mn(III, IV) to support metabolism or growth are classified into two groups. The first group use metals as their primary terminal electron acceptor for the oxidation of organic compounds, thereby conserving energy by Fe(III) or Mn(III, IV) respiration. The second fermenting group channel a portion of electrons to Fe(III) or Mn(IV) reduction, thus using these metals as nonrespiratory sinks (Lovley, 1987, 1997). The reduction of iron(III) and manganese(IV) by microbes is environmentally significant in coastal sediments and oxi- dizes organic matter to the range of 10%100% in aquatic sediments and submerged soils (Lovley, 2006). Most microbes that can reduce iron(III) can also reduce manganese(IV) although in general manganese oxides are Environmental variables and factors regulating microbial structure and functions 93 found in relatively low concentrations in coastal sediments when com- pared to ferric iron and sulfate. Nevertheless, like iron, reduced manga- nese can be reoxidized when diffusing upward to the oxic zone of the sediment. Because of the low abundance of the amorphous iron and man- ganese oxide minerals that can be readily utilized by bacteria, iron(III) and manganese(IV) reduction are generally not the predominant terminal elec- tron accepting process in coastal sediments. Bioturbation is an important agent of Fe(III) and Mn(IV) regeneration in coastal sediments (Gribsholt and Kristensen, 2002). The oxidation of both Mn(II) and Fe(II) is thermodynamically favored; however, the kinetics of the two processes are different. Mn(II) is kinetically stable, whereas oxidation of Fe(II) is very rapid and hence, biological catalysis is assumed to be unnecessary (Nealson, 1997). However, several lines of evidence prove that all iron reduction in nature is due to biological catalysis.

3.5.5.4 Sulfate reduction Below the zone of metal reduction, the major reduced species is sulfide, which is attributed to SRB and sulfate is the dominant electron acceptor in coastal sediments (Nealson, 1997; Vincent et al., 2017). Anoxic conditions favor the anaerobic process of sulfate reduction (Kristensen et al., 1994)inthe presence of high organic matter, where sulfate is reduced to sulfide by using electrons from low-molecular-weight organic compounds. A direct relation- ship exists between the abundance of SRB and sulfate reduction rates with depth. Although 80% of sulfate reduction occurs within the first six centi- meters of the sediment, diversity of SRB is the same throughout the depth of 22 the sediment. SO4 is one of the most significant and dominant environmen- tal factors in coastal sediments that explain the variations in bacterial commu- nity structure of SRB. Nevertheless, the diverse physiological abilities of SRB allow them to occur throughout the sediment, while their absolute abun- dance is limited by the quantity and quality of substrate available. Meanwhile, SRB play an important role in sulfur transformation through the process of dissimilatory sulfate reduction (Muyzer and Stams, 2008; He et al., 2015). This ultimately, mineralizes the organic material and increases the concentration of sulfide in the sediment (Kristensen et al., 1994), thereby exerting an important control on the pore water chemistry (Nickerson and Thibodeau, 1985).

3.5.5.5 Methanogenesis Methanogenesis is the final step in the anaerobic degradation of organic carbon. The important steps performed by methanogens are fermentation 94 Microbial Communities in Coastal Sediments

of acetate to CO2 and CH4 and oxidation of H2 to H2O. In the presence 2 22 of NO3 or SO4 , methanogens will be outcompeted by bacteria that respire more energy-yielding electron acceptors. For example, Fe(III) and Mn(IV) reducers interfere with the metabolism of methanogens by com- peting for organic carbon (Megonigal et al., 2004) and suppress methano- genesis as the iron and manganese reduction yields more free energy than methanogenesis. In addition, SRB are another group that interact closely with methanogens in the anaerobic pockets of sediments. The interaction is competitive for substrates, which depends on the presence of sulfate and salinity of the environment. Methanogenesis is influenced by various envi- ronmental factors and favored by highly reduced conditions (Reshmi et al., 2014).

3.5.6 Role of electron donors A major question is whether overall mineralization rates change when microbes use electron acceptors with relatively high free energy yield, 2 22 such as NO3 or SO4 , compared to methanogenesis. Among the various studies on the predominant TEA pathway through substrate amendments, some reveal greater rates of carbon mineralization in the presence of 2 22 NO3 compared to methanogenesis (Abell et al., 2009)orSO4 com- pared to methanogenesis (Pallud et al., 2007; Vincent et al., 2017). However, a difference in organic carbon mineralization rates under deni- trifying, sulfate-reducing, or methanogenic conditions was not observed by D’Angelo and Reddy (1999). The possible reason for this apparent dis- crepancy is attributed to the presence of electron donor (e.g., quantity and quality of organic carbon) that acts as a major factor governing the importance of TEAs in overall anaerobic mineralization. Hence, it can be conceived that the composition and activities of the microbial communi- ties are also regulated by the availability of carbon or electron donors (Torres et al., 2011). The rate of electron flow from the initial hydrolysis of organic com- pounds can be faster than electron consumption (Bruchert and Arnosti, 2003). In such situations dominated by an electron “imbalance,” the pres- ence of a more energetically favorable TEA may stimulate overall miner- alization. Contrastingly, where the electron source from organic matter is in balance with electron sinks from TEAs, overall mineralization rates may be less sensitive to the dominant TEA present. Moreover, the quality of carbon and the resultant differences in organic carbon mineralization Environmental variables and factors regulating microbial structure and functions 95 also control the electron donor supply to (Kelley et al., 1990; Schipper et al., 1994). Site differences in carbon quantity or quality were more important than TEA availability in regulating overall minerali- zation rates (Sutton-Grier et al., 2011). The compounds that support growth of methanogens as electron donors are H2, acetate, formate, alcohols and methylated compounds such as methanol, mono-, di-, and trimethylamines, and dimentyl sulfide (Jones, 1991). Among all substrates, H2 is the most important and nearly 73% of methanogenic species consume H2 (Megonigal et al., 2003). It is widely known that marine sediments are rich in methylated substrates, also referred as “noncompetitive substrates.” SRB compete with metha- nogens for substrates such as H2, acetate, and formate in the presence of 22 “ ” sufficient SO4 , which are referred as competitive substrates, but not for noncompetitive substrates (Capone and Kiene, 1988).

3.6 Biological factors The suite of factors influencing microbial community structure and ecol- ogy is complex owing to the fact that microbial community assembly is governed by a myriad of overlapping processes. Biological factors that influence microbial structure and activity in coastal sediments include fac- tors such as diversityenergy relationships, variable activity, behavior, and the various interactions between microorganisms involving trophic inter- actions such as competition, , and ; between microbes and plants as well as the influence of benthic animals in the sediments. In mangrove ecosystems, nitrogen fixation in sediments is associated with decomposing leaves in the rhizosphere (the layer of sediment close to roots associated with microorganismroot nutrient exchange) and the surface sediments have been shown to supply up to 60% of the nitrogen in sediments (Zuberer and Silver, 1979; Woitchik et al., 1997).

3.6.1 Trophic interactions Microorganisms are rarely encountered as single species population in the environment and hence, microbial interactions are crucial for a successful establishment and maintenance of a stable microbial population. Microbial associations are the result of a coevolution process that leads to adaptation and specialization, allowing the microbes to occupy different niches. This helps them to overcome various biotic and in the environ- ment as well as exchange of growth factors, genetic information, and 96 Microbial Communities in Coastal Sediments

Figure 3.5 Biotic factors influencing microbial community structure. molecular signaling (Braga et al., 2016). Coevolution of different species has resulted in a variety of relationship such as mutualistic, competitive, antagonistic, pathogenic, and parasitic relationships (Faust and Raes, 2012; Fig. 3.5). In coastal sediment, all basic types of bacterial interactions are encoun- tered, mostly between SRB and fermentative bacteria. When SRB grow on substrates produced by fermentative bacteria, the interaction is commen- salism (Fig. 3.5). By consumption of the metabolic products produced by fermentative bacteria to favor its growth, a mutualistic relation is observed. The sulfide produced by SRB represses growth of fermentative bacteria, which is amensalism. Competition is an important ecological mechanism that affects structure and function of these microbial communities in sedi- ments. For example, competition for substrates such as H2 and acetate is important, as they limit microbial activity in anaerobic sediments (Laanbroek et al., 1982; Megonigal et al., 2003). Contrastingly, exchange of H2 makes the microorganism’s metabolic partners in syntrophic rela- tionship and allows primary fermentation products to be completely min- eralized in anaerobic habitats. Such syntrophic relationships between two metabolically different bacteria depend on each other for energetic reasons.

3.6.1.1 Syntrophy and interspecies hydrogen transfer Syntrophy is very important in the anaerobic breakdown of organic mat- ter where two microbes cooperate to obtain a favorable amount of free energy released (Gibbs free energy, ΔG°) from an otherwise energetically unfavorable reaction for each. For example, secondary fermenting organ- isms utilize primary fermentation products that are kept energetically favorable through “interspecies hydrogen transfer” by anaerobic respirers Environmental variables and factors regulating microbial structure and functions 97 such as methanogens and sulfate reducers (Stams and Plugge, 2009). Accumulation of H2 may inhibit the ability of bacteria to ferment organic acids and alcohols by secondary fermentation, by causing the reaction to become endergonic. When the concentration of H2 increases above 2 10 4 atm, the secondary fermentation does not yield sufficient energy to support the growth of the fermenting bacteria. However, the H2 is con- tinuously consumed by syntrophic bacteria, which maintains the reaction as exergonic and such syntrophic relationship is a symbiotic association between two metabolically different bacteria that depend on each other for energetic reasons. Scavenging of H2 in syntrophic relationships is accomplished by a vari- ety of H2-utilizing anaerobes. For example, a large number of bacteria are capable of fermenting various amino acids syntrophically and the H2 formed is scavenged by methanogens. The process by which H2 is trans- ferred to the terminal organism is known as interspecies H2 transfer (Fig. 3.6; Boone et al., 1989; Thiele and Zeikus, 1988). The steps that occur after polymer hydrolysis in the breakdown of organic matter in anaerobic environments have been termed “intermediary ecosystem

Figure 3.6 Syntrophism: interspecies H2 transfer. 98 Microbial Communities in Coastal Sediments metabolism” (Drake et al., 2009) where syntrophic fatty and aromatic acid metabolism accounts for much of the carbon flux in methanogenic envir- onments (McInerney et al., 2009). Predation by protozoa and parasitism by virus contribute to selective loss of certain microbial communities in the sediments, although the movement of protozoa may be restricted by the complexity of sediments (Torsvik et al., 2002). Viruses are abundant in sediment pore water and parasitism by host-specific viruses will allow coexistence of different bacterial taxa within the bacterial community, by controlling the high growth rate of certain bacterial species by way of “killing the winning population.” Ironically bacterial diversity determines the abundance of viruses present by means of the differences in the growth rate between coexisting bacteria (Thingstad, 2000). In coastal sediments, protozoan predation is largely restricted, unlike the water column, which explains the large observed differences in total bacterial abundance.Thisisbecausethecomplexmatrix of soils and sediments acts as an obstacle to the movement of protozoa than to the diffusion of small viruses. Moreover, viruses are abundant in sediment pore water (Drake et al., 1998). Bacterial diversity is controlled by viral lysis by allowing competing bacterial species to coexist. For example, virus- induced bacterial cell death is a significant contributor to mortality of sediment bacteria (Danovaro et al., 2008). Interestingly, viral lysis directly affects microbial community composition (Weitz and Wilhelm, 2012), and indirectly influences the ecosystem function by means of releasing organic carbon and nutrients by a process known as “viral shunt” (Fig. 3.7). The viral shunt is responsible for releasing 0.370.63 gigatons of carbon per year on a global scale, in the form of labile organic matter, utilized by prokaryotes only (Danovaro et al., 2008).

3.6.2 Evolutionary mechanisms and diversification Interactions between the ecological factors and the intrinsic evolutionary mechanisms also exhibit control on microbial diversity. One reason for the high genomic diversity observed in sediment microbial communities is the large populations of organisms and the capacity to accumulate large numbers of mutations. Each population represents a mixture of genetically diverging clonal cell lines on which natural selection acts. Molecular mechanisms such as lateral DNA transfer and recombination, are also facil- itated by high population densities, and may influence genetic diversity. If lateral transfer occurs within a group of closely related bacteria, it will Environmental variables and factors regulating microbial structure and functions 99

Figure 3.7 Viral shunt and microbial community composition. ensure genetic coherence and slow diversification. By contrast, gene trans- fer and recombination across species and genus barriers could promote environmental adaptation and the evolution of new traits (e.g., the trans- fer of antibiotic resistance among different species of bacterial pathogens), thereby increasing diversity (Torsvik et al., 2002). Random changes in species-relative abundances due to stochastic vari- ables, such as random birth 2 death events within a population or unpredictable disturbances are referred as ecological drift (Vellend, 2010; Nemergut et al., 2013). This concept is analogous to that of genetic drift in population genetics, whereby changes in gene frequencies in a popula- tion occur solely by chance (Chase and Myers, 2011). Although the fac- tors influencing drift are largely unpredictable, they can still have significant effects on the composition of microbial communities (Stegen et al., 2012, 2013). The generation of new genetic variation due to spontaneous muta- tions, gene loss and genome rearrangements, or horizontal gene transfer (HGT) is defined as diversification (Prozorov, 2001; Albalat and Cañestro, 2016). These processes hinge on community growth or a succession of multiple generations for genetic changes to settle in a population 100 Microbial Communities in Coastal Sediments

(Weller and Wu, 2015). High rates of HGT occur in dynamic environ- ments with higher growth rates of bacteria as well as mutations that intro- duce novel traits into populations and drive speciation events. In contrast, slow-growing communities have little scope for adaptive change. Hence, energy limitations in sediments pose constraints on diversification by these slow-growing communities (Lever et al., 2015; Jørgensen and Marshall, 2016). This in turn reduces growth rate of microbial biomass, which serves as a proxy for the generation time of microbial populations.

3.6.3 Ecological coherence Selection of individuals or a population in a microbial community is imposed by abiotic and biotic pressures within the sediment environment, the latter of which can be either antagonistic or synergistic (Stegen et al., 2012). Selection is a mechanism of community assembly due to fitness dif- ferences between taxa (Nemergut et al., 2013) and the effects of selection typically manifest in the phylogenetic clustering of communities within similar environments. This is believed to be the result of ecological coherence within taxa, which suggests that phylogenetic units have ecological impor- tance (Philippot et al., 2009). Ecological coherence has been demonstrated at high taxonomic levels, resulting in assemblages of microorganisms at the phyla or class level becoming clustered within similar environments (Philippot et al., 2009). Nevertheless, competition for resources can alter this picture, whereby species occupying a single habitat can be more dis- tantly related than expected by chance resulting in a pattern of phyloge- netic overdispersion (Koeppel and Wu, 2014). The effects of selection are similar to those of diversification so that both produce phylogenetic or functional patterns in response to environ- mental parameters. Diversification introduces new populations into a community, whereas selection only affects the relative abundances of populations that are already present (Nemergut et al., 2013). This reveals the fact that persisting taxonomic groups possess diverse metabolisms, with discoveries of new metabolic capabilities increasing dramatically in recent years (Fullerton and Moyer, 2016; Lazar et al., 2016; Nobu et al., 2016). These metabolisms could confer an advantage in the sediment, as they would provide energy regardless of changing geochemistry. For example, there are many groups that are capable of growing fermentatively or by acetogenesis in deeper sediment layers. This is in accordance with the the- ory proposed by Lever et al. (2010), who suggested that the wide substrate Environmental variables and factors regulating microbial structure and functions 101 spectrum available to acetogens may allow them to outcompete substrate specialists under extreme resource limitation. Functional genes involved in fermentation greatly outnumber those responsible for terminal oxidation processes (Kirchman et al., 2014). In addition to selection imposed by metabolic capabilities, functional differences among taxa may also represent selection pressures other than, or in addition to, energy limitation. Both viral-induced and other forms of mortality are likely to contribute to the selective growth of surviving populations. The most notable of these is viral-induced cell death that has been suggested to be a significant contributor to cellular mortality in the deep subsurface (Danovaro et al., 2008; Engelhardt et al., 2015; Jørgensen and Marshall, 2016). Resistance to viral lysis could mediate the selective survival of key taxa in the sediment column while also playing an impor- tant role in altering the subsurface environment. However, mortality can only account for a minor fraction of the organic carbon assimilated (Lomstein et al., 2012) and unadapted populations that die off further increase the selective growth of persisting taxa.

3.6.4 Microbial characteristics The quality and quantity of organic matter determine the energy flux available in the sediment during a particular time and the relationship between microbial community size and energy availability is believed to depend on basal power requirement (BPR) that refers to the minimum amount of energy required for basic repair and maintenance functions of microbial cells (Lever et al., 2015). BPR determines the theoretical upper limit of the community size that can be supported in the available energy flux (Hoehler and Jørgensen, 2013). Active cells may, however, exist in a survival state, devoting their limited energy to the repair and synthesis of essential biomolecules rather than to cell division and growth (Kempes et al., 2017). Microorganisms that lack such adaptation may persist throughout the sediment column by sporulation that allows them to switch to a dormant state when the energy flux drops below their BPR (Petro et al., 2017). Response of microbial communities is also related to the catabolic diversity of soil microorganisms. Catabolic response profile has been widely applied as a method to characterize the microbial functional diver- sity. A greater catabolic response to a substrate in one system as compared with another indicates that the microbial community is more functionally 102 Microbial Communities in Coastal Sediments adept in utilizing that substrate and may indicate previous exposure to the associated carbon sources (Baldock et al., 2004; Degens and Harris, 1997). In addition, active dispersal in the form of motility can allow bacteria to retain access to freshly deposited sources of organic matter over longer timescales. Nevertheless, motility usually results from chemotactic attrac- tion to low-molecular-weight organics or gradients or terminal electron acceptors (Fenchel, 2008). In sediments, motile capabilities are largely par- adoxical. They confer selective advantages in that they provide faster and directional access to energy sources within a highly energy-depleted envi- ronment. However, they do so at a high energetic cost, due to the need for maintenance and operation of specialized structures, such as the flagella (Hoehler and Jørgensen, 2013; Taylor and Stocker, 2012). While there has been limited research on the motility of bacteria in deeper sediment layers, it can be understood that the low energy and homogeneity of the environment would not support a motile lifestyle (Lever et al., 2015). The differences in microbial community structure and diversity are a manifestation of ecophysiological requirements of the bacteria, in particu- lar with regard to the availability of DOC substrates. DOC is labile to degradation in the surface sediment and becomes increasingly resistant to degradation in deeper sediments, probably due to the formation of com- plex polysaccharides and polysulfides. Microorganisms adapt to exploit the various pools of organic carbon throughout the sediment depth indicating a vertical niche separation. For example, few phylotypes such as alpha-, beta-, and deltaproteobacteria are present only in the top and middle depths and are absent in the bottom (Torsvik et al., 2002). This reflects the specialization of certain phylotypes to specific ecological niches within the sediment that is due to a difference in the quality of OM. Thus the partitioning of resources within the sediment creates specific niches, enhancing microbial specialization and division into distinct eco- logical guilds. Change in phylogenetic richness with depth occurs due to the fact that the availability of electron acceptors and energy sources decrease with depth. The abundance of active bacteria is directly related to the availability of electron acceptors and energy sources within the sed- iment and decreases with depth. Due to increasing substrate limitation with depth, DOC becomes unavailable to the microorganisms with depth. This, in turn, results in vertical depth gradients of organic compounds that serve as energy sources for microorganisms. The addition of phytodetritus to marine estuarine sediment microcosms stimulates microbial activity leading to faster rates of oxygen consumption. This suggests that the high Environmental variables and factors regulating microbial structure and functions 103 phytodetritus-loaded coastal sediments favored the establishment of anaer- obic microorganisms such as SRB and MA.

3.6.5 Bioturbation and ventilation Animals living in the coastal sediments modify the texture and geo- chemistry of the sediments by their physical movements that allow particle mixing, enhanced transport of oxygen and nutrients into the sedi- ments, which are referred to bioirrigation/bioturbation/bioventilation (Middelburg and Levin, 2009). Active infaunal ventilation of burrows and tubes by benthic macroinvertebrates by sediment reworking is a major factor controlling biogeochemical processes occurring in sediments (Kristensen, 1984, 1985), which in turn influence the structure and func- tion of sediment microbial communities. Thus ventilation activity results in renewal of burrow water, which aids in gaseous exchange as well as food transport and produces an increased transport of water and solutes in and out of the sediment. Meiofauna in the sediments facilitates biominer- alization of organic material and enhances nutrient regeneration indirectly by stimulating microbial community through release of excretion products as well as bioturbation (Coull, 1999; De Troch et al., 2005). Small-scale biological interactions between bacteria, copepods, and diatoms can have an important impact on denitrification and also on sediment nitrogen fluxes (Stock et al., 2014). The supply of oxygen in burrows is primarily dependent on the venti- lation activity of the burrow inhabitants. Under nonventilation conditions, the available oxygen is consumed in the surface layer of coastal sediments, which diffuses to a depth of only a few millimeters (Jørgensen and Revsbech, 1985). Nevertheless, majority of macroinvertebrates perform an intermittent ventilation pattern, which promote variable oxygen condi- tions in the burrows. This is explained in case of nereid polychaetes that have long resting periods followed by active ventilation periods. The oxy- gen level is almost similar to surface water during active ventilation peri- ods; however, during resting periods, oxygen consumption by the burrow inhabitant including wall microbes rapidly exhaust the oxygen. Such con- ditions may steepen the gradients of microbial processes and solutes. Degree of oxygenation differs with species, for example, the oxic sedi- ment volume is 30%50% with the polychaete Nereis virens (Hylleberg and Henriksen, 1980), and 100%150% with the crustacean Corophium volutator. Oxygen availability in burrow environments is of vital 104 Microbial Communities in Coastal Sediments importance for the macrofaunal inhabitants, but it may also affect growth and population sizes of the associated microorganisms, which are normally quantitatively and qualitatively different from both the ambient anoxic and the oxic surface sediment. Nitrification and denitrification are the major microbial activities that occur in the burrows of infaunal animals. Penetration of gases such as oxygen into sediments is important for such microbial nitrogen transfor- mations (Hansen et al., 1981; Andersen et al., 1984). Ammonium, the precursor for nitrification, present in burrow walls is oxidized by autotro- phic nitrifying bacteria or passed to the overlying water by ventilation current. Nitrification in burrows of a natural population of N. virens accounts for 10%70% (average 40%) of the bulk sediment nitrification Kristensen (1985). The nitrate produced by nitrification in the oxic layers will either pass to the overlying water or enter the anoxic sediment where it is reduced to free nitrogen gas by denitrification, and thereby lost from the ecosystem (Vanderborght and Billen, 1975). Denitrification is primar- ily dependent upon certain favorable factors such as anoxic conditions, availability of nitrate, and the presence of energy source that are appar- ently satisfied in the mucus-lined burrows of benthic animals, especially during periods of ventilatory rest. Most of the nitrate produced in the burrows is reduced by denitrifiers compared to the surface sediment, where only 6%35% of the nitrate produced generally is consumed by denitrification (Billen, 1978; Nishio et al., 1982).

3.6.6 Plant interactions The relationship between sediment microbes and plant communities is mutualistic, with plants oxygenating and enriching the sediments with organic matter, and microbes, in turn, degrading organic matter thus releasing nutrients for plant uptake. Hence, structure and function of sedi- ment microbial communities are shaped based on the plant detritus including leaves and root exudates that constitute the major carbon and organic matter input in the sediments. This is particularly evident in the ecosystems such as mangroves (Holguin et al., 2001), where mineraliza- tion of organic matter from mangrove root exudates, leaves, and phyto- plankton detritus in mangrove sediments occurs through the action of several groups of microorganisms. Oxygen is a principal electron acceptor, which is released from mangrove roots and rapidly consumed during res- piration of microorganisms (Canfield, 1993) highlighting the importance Environmental variables and factors regulating microbial structure and functions 105 of living mangrove trees for nutrient mineralization in mangrove sedi- ment. The roots, in turn, benefit from the remineralization of the organic matter and nutrients by the microorganisms (Bashan and Holguin, 1997; Bashan et al., 2000). Plants in wetlands promote aerobic metabolism in sediments by sup- plying O2 to their root system, where some of the O2 leaks into soil and penetrates the sediments for a few millimeters. Plants are also a source of labile organic carbon compounds that fuel anaerobic metabolism (Megonigal et al., 2003). Conversely, microbial communities also posi- tively influence growth of plants such as seagrasses, for example, which are highly productive in oligotrophic coastal ecosystems. The functional capacity of microbial communities contributes to the high productivity of seagrass meadows under such oligotrophic conditions. This was evident from the functional diversity of microbial communities in the seagrass ecosystem sediments, which showed a high abundance of phosphorus and sulfur metabolism genes (Fraser et al., 2018). Such key functional gene abundances are recognized as drivers that influence the productivity and structure of seagrass ecosystems.

3.7 Nutritional factors Coastal habitats with constantly low levels of nutrients represent a major and important extreme environment. Survival of microbial communities in such nutritionally deprived environment must involve the expression of genes that has evolved over long periods of good and bad environmental conditions. In this context, understanding the contrast between oligo- trophs and the copiotrophs and their ways of life is extremely important for the process of microbial evolution. Some types of organisms are usu- ally found in environments with low levels of nutrients and are labeled “oligotrophs.” In concept, the class of bacteria live in a peren- nially sparse environment and they have an allochthonous mode of feed- ing (Koch, 2001). They exhibit K-strategist life histories with a low growth rate (Fierer et al., 2007). Copiotrophs, or eutrophs, are associated with richer environments and are generally adapted to using a resource rapidly when available. Such microbes are “zymogenous,” which means they have the ability to ferment carbohydrates (Koch, 2001). These microorganisms will exhibit r-strategists life histories, and have a high growth rate (Fierer et al., 2007). 106 Microbial Communities in Coastal Sediments

Bacteroidetes, which typically contains bacteria with hydrolytic and fer- menting abilities (Weller et al., 2000) can be classified as copiotrophs and are dominant in environments with high carbon availability and reminera- lization rates (Fierer et al., 2007). Although both oligotrophs and copio- trophs can survive in a poor nutritional environment, only oligotrophs persist in chronic starvation conditions and, conversely, may not be able to persist for long periods in richer environments (Koch, 2001). Nutrient enrichment or eutrophication also leads to responses in microbial diversity. An increase in nutrient supply promotes domination of few opportunistic species, decreasing the evenness of species distribution (Torsvik et al., 2002).

3.8 Natural and anthropogenic disturbances Environmental disturbances will ensure that communities include a mix- ture of different stages of succession. Sediments of disturbed areas contain fewer dominant bacterial species that is characteristic of disturbed and harsh environments (Atlas et al., 1991). However, strong and frequent dis- turbances will cause the disintegration of the microhabitats and disruption of the boundaries between populations, allowing local resources to become available to a larger proportion of the entire microbial biomass. Dredging of sediments significantly alters the composition and structures of sediment benthic microbial communities (Zhang et al., 2017). Nevertheless, undisturbed sediments are characterized by higher bacterial diversity that provides more diverse functional pathways, which play an important role in ecosystem functioning (Levin et al., 2001). Although disturbance may create empty niches by killing or inactiva- tion of microbial communities, dispersal can facilitate dissemination, thus altering the community composition (Zhang et al., 2017). In a study con- ducted to establish relationship between soil properties and relative abun- dance of various bacterial phyla in anthropogenically disturbed reclaimed wetlands and undisturbed natural wetlands, it was observed that the rela- tive abundance of Deltaproteobacteria decreased in disturbed sediments, which was attributed to a decrease in TOC concentration in disturbed wetlands. In contrast, the relative abundance of Gammaproteobacteria was higher in disturbed than natural wetlands because the disturbed wetland had significantly higher N (Campbell et al., 2010; Nacke et al., 2011). The more undisturbed sediments, with higher phosphate and nitrogen content may favor nitrogen-fixing bacteria (Sjoling et al., 2005). Environmental variables and factors regulating microbial structure and functions 107

3.9 Presence of contaminants/toxic substances Both the water and sediments act as reservoirs for pollutants from land- based sources via river runoffs and sewage outfalls, such as nutrients, poly- cyclic aromatic hydrocarbons, pesticides and heavy metals (Lin et al., 2002; Liu et al., 2001, 2008). Such pollutants in water can be diluted or removed by tides. On the contrary, pollutants in sediments can exist for a relatively long time, inevitably affecting the sediment-borne microbial communities and their functions (Wang et al., 2016). In extreme cases, an accumulation of toxic metabolites or other detrimental effects can occur, and are likely to reduce diversity even more. Therefore investigation of the diversity and abundance of microorganisms, especially of bacteria and their applications as a is becoming increasingly important to predict the environmental changes. Bacterial diversity can be lower in stressed environments, for example, in areas prone to heavy metals (Hu et al., 2007) or hydrocarbon contami- nation (Greer, 2010). Nevertheless, availability of a wide variety of sub- strates could result in high taxonomic and metabolic diversity. The levels of pollutants are not a deterrent for affecting bacterial diversity and con- trastingly, they could play an important role in buffering the anthropo- genic influences. The prevailing environmental conditions could be crucial in influencing the composition of the autochthonous benthic bac- terial communities, which play a vital role in ecosystem functioning. Langenheder and Prosser (2008) used RNA stable isotope probing to investigate how the diversity and community structure of a heterotrophic benzoate degrading bacterial community is influenced by resource con- centration. They found that the composition of the benzoate-degrading community changed with benzoate concentration with a decrease in taxa evenness at higher concentrations. There were generalists active at all resource concentrations and specialists active at one particular concentra- tion or low and high concentrations. Interestingly, a significant fraction of the novel bacterial population in stressed environments await to be cul- tured and identified for various applications (Fernandes et al., 2014). Heavy organic input leads to a decrease, not necessarily in the number of species present, but in the evenness of species distribution of the com- munity (Torsvik et al., 2002). In Jiaozhou Bay of China, the presence of heavy metals in the sediment was found to influence the bacterial com- munity structure (Yao et al., 2017). In the intertidal sediments of Yangtze estuary, China, the bacterial community structure was affected by river 108 Microbial Communities in Coastal Sediments runoff and sewerage discharge; particularly, sulfate, salinity, and total phosphorus were the environmental variables that influenced the commu- nity structure (Guo et al., 2018). Contrastingly, microbial communities in the sediments are also reported to remove the contaminants by degradation (discussed in detail in Chapter 5: Role of Microbes in Biodegradation and Biotransformation of Persistent Organic Pollutants). Hydrocarbon-degrading community was found to be well established in an oil-contaminated mangrove sediment, with more abundance in the top layer (Andrade et al., 2012). Higher bac- terial diversity and richness at the anthropogenically influenced Divar than pristine Tuvem, two mangrove ecosystems in Goa, India suggest that the levels of pollutants are not deterrent for affecting bacterial diversity, which means the pollutants are within the level to maintain the microbial diver- sity. Hence, it can be concluded that the prevailing environmental condi- tions could be crucial in influencing the composition of the benthic bacterial communities, which play a vital role in ecosystem functioning (Fernandes et al., 2014).

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4.1 Introduction Human activities such as changes in land use and land cover, fertilizer appli- cation, and wastewater discharge have increased eutrophication of coastal ecosystems, thus affecting their biogeochemistry (Jennerjahn et al., 2008; Smith et al., 2006). Eutrophication leads to excess accumulation of nutrients such as carbon, phosphorus, and nitrogen in coastal sediments. Primary pro- duction is generally limited by the availability of phosphorus and nitrogen, and is therefore enhanced by the external nutrient load. Nutrient loading and associated higher rates of primary production enhance the production of organic matter (OM) and sedimentation (Boström et al., 1988). The elevated amount of OM in sediments increases decomposition, O2 consumption, and the production of carbon gases such as carbon dioxide (CO2)andmethane (CH4). Mineralization also liberates nutrients from the OM back into the water column where they can be used again in primary production. This internal loading of nutrients is an underlying mechanism in maintaining long-term eutrophication of lakes. Human-caused disturbances on coastal wetland release carbon from sediments, turning them into a strong net source of greenhouse gas (GHG) emissions, irrespective of their GHG balance in the natural state. Some coastal wetlands emit CH4, a GHG 25 times more potent than CO2. Methane production is generally more intense in brackish and fresh- water tidal flats and marshes because of the high OM content of the soils at anoxic depth. Another GHG of concern in coastal environment is nitrous oxide (N2O). Nitrous oxide is mainly formed as a by-product dur- ing nitrification and as an intermediate during denitrification. Although ammonia and nitrate are natural constituents in coastal ecosystems, they are now found at heightened levels due to runoff from intensive agricul- ture and other anthropogenic sources such as air pollution. Ongoing deg- radation of coastal ecosystems and associated emissions of GHGs, as well

Microbial Communities in Coastal Sediments © 2021 Elsevier Inc. DOI: https://doi.org/10.1016/B978-0-12-815165-5.00004-2 All rights reserved. 119 120 Microbial Communities in Coastal Sediments as lost sequestration are currently neither recognized as a significant driver of climate change, nor mitigated.

4.2 Biogeocycling of nutrients Biogeocycling of nutrients in coastal sediments is linked to utilization of certain nutrients as electron donors as well as electron acceptors during their respiration process. The OM in sediments is assimilated by hetero- trophic microorganisms or respired and mineralized. The conceptual model of terminal electron acceptors (TEAs) by Billen et al. (1980) throws light on the consumption of nutrients during respiration in sediments. The electron acceptors preferentially used according to their free energy are utilized by the respective microorganisms in the sequence as follows: 2 22 O2NO3 Mn(IV)Fe(III)SO4 fermentation (Zehnder and Stumm, 1988). Oxygen as the TEA occurs under oxic conditions; nitrate, Mn, and Fe under suboxic conditions; and sulfate reduction as well as methanogenesis under anoxic environments (Froelich et al., 1979). Both natural and anthro- pogenic influenced processes govern the oxygen balance and dissolved oxy- gen dynamics in coastal sediments (Middelburg and Levin, 2009). Although the first postulated conceptual model shows distinct layer-wise occurrence of the various electron acceptors, several evidences report the terminal electron processes such as sulfate reduction, denitrification, and methanogenesis occurring in the same horizons (Lovely and Philips, 1987). This may be explained by the presence of microhabitats (microniches), that is, spaces of more reduced conditions within a spatial structure in the sediment layers (Jørgensen, 1977). Anaerobic mineralization of OM results in the formation of various reduced substances such as ammonium, iron(II), manganese(II), hydrogen sulfide, and methane (Middelburg and Levin, 2009). Changes in sedimentary oxygen level consequently change the diagenetic pathways, where hypoxic conditions may lead to less degradation of OM. Hence, oxy- gen availability in the sediments can be considered as one of the main factors governing diagenetic pathways as well as sediment biogeochemistry.

4.2.1 Carbon Coastal ecosystems are the most productive as they receive allochthonous input of OM and hence, the primary productivity is generally greater than in open oceans (Malone et al., 1996). Nevertheless, these ecosystems are considered as sinks and may act as traps for terrestrial OM. Human intervention within watersheds has led to increase in riverine discharges. Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 121

Net nutrient fluxes in the coastal zone can be determined from budget calculations, which are essential to evaluate the effects of riverine dis- charges on coastal function and carbon metabolism. The magnitude of removal or addition of OM in coastal ecosystems is seasonally variable, and these processes must be considered for net flux estimates from the river to the sea (Hung and Huang, 2005). Radioisotopic studies reveal that the OM discharged is composed of two groups: one of them is of land plant origin characterized by lower δ13C, which tends to deposit in the coastal sediments, and the other one is characterized by highly degraded soil-derived material with higher δ13C, which can be further transported offshore (Gordon and Goñi, 2003). The terrestrial organic car- bon supports the metabolism of heterotrophic bacteria in the water col- umn thus increasing the amount of dissolved organic carbon (DOC) (Jansson et al., 2006). Approximately 30% of the organic carbon buried in the coastal and marine sediments is of terrestrial origin. Of the 0.2 Pg 2 C year 1 discharged by rivers into coastal ecosystems, nearly 50% is decomposed by microorganisms comprising of primary degraders that ini- tiate digestion of woody and nonwoody plant materials followed by sec- ondary degraders that consume the metabolites from primary degraders 2 (Bianchi, 2011). The remaining 50% (0.1 Pg C year 1) is buried in the sediments. Fig. 4.1 schematically represents the quantity of organic carbon reaching the coastal sediment from various terrestrial sources and burial in the sediments. Moreover, the photosynthesizing microorganisms in the epipelagic zone also contribute to autochthonous organic carbon pool in the coastal ecosystems. Most part of this net primary production is funneled through dissolved organic matter (DOM) into the microbial loop, where respira- tion converts a substantial part of the primary production to dissolved inorganic carbon (DIC) by respiration. Freshly produced DOM is con- sumed in the water column within minutes or days of production, thus preventing its accumulation or export to the sediments (Dittmar and Stubbins, 2014). Fig. 4.2 depicts the pools of organic carbon (POC, DOC) and DIC in the water column and sediment of oceans. In the sediments, OM, an important source of organic carbon, is min- eralized through heterotrophic decomposition (Megonigal et al., 2004) and carbon cycle is an energy cycle for the sediment microbes. Decomposition often begins with physical fragmentation, proceeds to exoenzyme-mediated hydrolysis of complex organic compounds, and ends with mineralization of simple organic compounds to gases via 122 Microbial Communities in Coastal Sediments

Figure 4.1 Global carbon cycle (Bianchi, 2011). microbial respiration. This last step in the overall decomposition process ultimately requires the flow of electrons from OM (electron donors) to one of the several TEAs (Müller et al., 2005). In eutrophic systems, OM decomposition is high and oxygen is depleted rapidly in the sediment dur- ing aerobic degradation. Further degradation is carried out under faculta- tive and obligate anaerobic conditions. Here complete oxidation is achieved by sequential action of different groups of anaerobic bacteria (Capone and Kiene, 1988). Initially, the cellulolytic bacteria hydrolyze the polymers to monomers and further breakdown the monomers to alcohols, fatty acids, and hydrogen. Alcohols and fatty acids are further degraded by syntrophic bacteria to acetate, H2, and CO2, which are fur- ther utilized by methanogens (Zinder, 1993; Conrad, 1999). Microbes mediate the conversion of inorganic to organic carbon by autotrophic process and the reciprocal conversion of organic to inorganic carbon by heterotrophic process. Autotrophic process involves various elec- tron donors such as H2O, H2S, and Fe(II) and heterotrophic process involves various electron acceptors (Fig. 4.3). Inorganic carbon is reduced to organic form through various biochemical pathways, thus, changing the oxidation state of carbon from 1 IV to states ranging from 1 III to IV. Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 123

Figure 4.2 Pools and flux of POC/DOC in the oceans (Dittmar and Stubbins, 2014).

Certain coastal ecosystems such as mangrove forests and seagrass mea- dows take up substantial amount of carbon by photosynthesis and store a significant fraction of that carbon in the anaerobic sediments (Chmura et al., 2003; Donato et al., 2011; Fourqurean et al., 2012). This carbon has been termed as “blue carbon.” Such coastal blue carbon ecosystems 124 Microbial Communities in Coastal Sediments

Figure 4.3 Microbe-mediated carbon cycle (Megonigal et al., 2004). are some of the most carbon-rich ecosystems in the planet (Mcleod et al., 2011). This is a major reason for efforts taken to include coastal ecosys- tems in international climate protection activities and policy frameworks (Wylie et al., 2016; Howard et al., 2017). Vegetated coastal ecosystems contribute to 0.2% of area of ocean, but have stored blue carbon stock equivalent to 50% of carbon in the sediments (Duarte et al., 2013). Disturbing coastal hydrology can enable oxygen to oxidize stored OM in sediments, causing the release of CO2 and CH4, and transforming the coastal ecosystems as source of GHGs (Pendleton et al., 2012). Future global changes, particularly those related to climate change impacts, global warming, and ocean acidification using modeling studies predict shift in the microbial community structure in coastal ecosystems. Warming stimulates harmful algal blooms, in addition to significant changes relating to shift from large cells such as diatoms to smaller pico- cyanobacteria. This, in turn may influence the oceans’ biological pump, wherein, the reduced cell sizes lead to reduced storage of sinking POC (Hutchins and Fu, 2017). Increasing levels of anthropogenic CO2 entering the oceans will increase the size of DIC pool resulting in ocean acidifica- tion. Although the planktonic assemblages have shown a shift in taxo- nomic dominance in relation to effects of future ocean acidification, little or no effects were observed with regard to heterotrophic bacterial popula- tion and community structure. Future projected changes include increased CO2 uptake by seawater and lowering ocean pH resulting in reduced photosynthesis and restricted supplies of organic carbon to the sediments. Reduced release of CO2 through microbial respiration results in reduced Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 125

Figure 4.4 (A) Current and (B) future predicted changes in oceanic carbon cycle (Hutchins and Fu, 2017). Blue arrows indicate processes that result in uptake of DIC; green arrows indicate processes that release CO2. vertical export flux of organic carbon and hence its reduced flow in the (Fig. 4.4). Ultimately, the downsized biological carbon cycle reduces the capacity to store CO2. This, in turn affects the coastal biodiversity and the supply of harvestable biological resources.

4.2.2 Nitrogen Microbial mediated nitrogen cycle includes multiple processes, and ele- vated sediment OM content and high content of N in OM increase the release of nitrogen from sediments. Nitrogen is liberated from the sedi- 1 ments as particulate N, dissolved organic N, ammonia (NH4 ), nitrate (NO3), nitrous oxide (N2O), or N2 as a result of OM mineralization, nitrification, and denitrification (Keeney, 1973)(Fig. 4.5). Bacteria have unique roles and participate in all the important nitrogen transformation processes in the sediments (Megonigal et al., 2004). The proportion of various N compounds released from the sediments is dependent on O2 availability and the amount of OM (Keeney, 1973; van Lujin et al., 1999). Penetration of O2 is important for microbial nitrogen transforma- tions such as nitrification and denitrification (Andersen et al., 1984; Sørensen, 1984) in the sediments. The mineralization of OM produces 1 NH4 liberating organic N bound in proteins and nucleic acids to 1 NH4 . With decline in oxygen concentrations, significant increase in ammonium effluxes can be observed (Middelburg and Levin, 2009). The 126 Microbial Communities in Coastal Sediments

Figure 4.5 Microbe-mediated nitrogen cycle (Megonigal et al., 2004). net rate of N mineralization in sediments is equivalent to the rate of 1 1 NH4 production minus the rate of NH4 incorporation into cells 1 (Blackburn, 1979). The possible fates generally considered for the NH4 (1) that is produced in the sediments and (2) that is not incorporated into 1 cells are as follows: (a) enters the sediment NH4 pool, either in pore- water solution or adsorbed to particles; (b) transported to the overlying 1 water. Under anoxic conditions, NH4 is released into the overlying 1 water, whereas, in the presence of O2, some of the NH4 is oxidized 2 during nitrification to NO3 . Nitrate is either diffused upward to the overlying water or downward to more reduced sediments, where it is reduced in denitrification to N2O and/or N2 and thereby lost from the ecosystem. It is estimated that nearly 27%57% of the reduced nitrate is accounted for by denitrification and the remainder is reduced to ammo- nium by dissimilatory pathways (Koike and Hattori, 1978). Liberation of gases such as N2O and N2 is significant in controlling eutrophication by removing nitrogen from a coastal ecosystem into the atmosphere 1 2 (Seitzinger, 1988). The dissolved organic N, NH4 and NO3 , which can be used in primary production, is, on the other hand, of most impor- tance in maintaining eutrophication. The loss of nitrogen via denitrifica- tion includes the input of nitrogen via nitrogen fixation in coastal sediments. Denitrification removes an amount of nitrogen equivalent to 20% and 50% of the external nitrogen input in estuaries (Seitzinger, Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 127

1988). Denitrification is primarily dependent upon anoxic conditions, availability of nitrate, and the presence of energy source. Relationship between denitrification and nitrate may vary with depth of the sediment (Kaspar, 1983) and also spatially (Seitzinger, 1988). Whatever the source of nitrogen, either by denitrification or from overlying water, a positive correlation exists between denitrification rate and nitrate concentration. From water column, nitrate diffuses into the sediment by bioturbation, which is again denitrified. Hence, the exchange of nitrate between sedi- ment and overlying water is affected considerably by the burrow-dwelling infauna. Although less, this might act as a major source of nitrate for deni- trification in the sediment. In the presence of different infaunal species, the flux of nitrate either increases or decreases. This may be attributed to substantial variation of the coupling and rates of nitrification and denitrifica- tion activity in the burrow environments (Kristensen, 1988). Nitrification and denitrification are spatially separated as nitrification occurs under aerobic or microaerobic conditions and denitrification occurs under anaerobic condi- 2 tions. However, NO3 released during nitrification can be consumed rap- idly by coupled nitrificationdenitrification (Megonigal et al., 2004). In coastal sediments, denitrification can be uncoupled from nitrification when 2 there is substantial supply of NO3 from outside sources such as fertilizers, atmospheric deposition, and terrestrial runoff. In such cases, sedimentary 2 denitrification rates will be directly proportional to the NO3 in the water 2 column, where NO3 diffuses into the sediment. Coupled nitrifica- 1 tiondenitrification occurs in the presence of NH4 and organic carbon, 2 when the input of NO3 from external sources is limited and nitrification fuels denitrification in the sediments (Salahudeen et al., 2018). Removal of N2 from coastal ecosystems, which counteracts eutrophi- cation, is limited by the availability of oxygen. When anoxia develops, nitrification ceases and removal of N2 cannot proceed due to lack of nitrate. Denitrification occurs in thin unstable layers of the sediment, where oxygen is sufficiently low or completely depleted and nitrate is still present. Here, the heterotrophic bacteria shift to utilize nitrate as a TEA. 2 Denitrifying enzymes in the increasing order of oxidation state are NO3 2 reductase, NO2 reductase, NO reductase, and N2O reductase. These enzymes are sensitive to H2SandO2 and inhibited in an order that is inversely proportional to the oxidation state of nitrogen. Chemolithotrophic denitrifying bacterial communities found in deeper layers of coastal waters 2 reduce NO3 to N2 gas by using H2Sasenergysource.Apartfromdenitri- fication, dissimilatory nitrate reduction to ammonium (DNRA) is less 128 Microbial Communities in Coastal Sediments

2 explored sink of NO3 . Both these processes compete for electron donor 2 (carbon) and electron acceptor (NO3 ) and although they can occur simul- taneously, DNRA has a competitive advantage. However, when organic carbon is limiting, denitrification outcompetes DNRA. Moreover, there are several lines of evidence that support that more labile organic carbon and high temperatures favor DNRA bacteria in coastal sediments (King and Nedwell, 1984, 1985). DNRA coupled to nitrification has been reported in 2 Gulf of Finland at an oxygen concentration of 3.4 mg L 1 (Jäntti and Hietanen, 2012). Although DNRA was found to dominate at higher oxy- gen conditions in the seasonally hypoxic coastal areas of Southern Baltic Sea, it was not observed in the northern Baltic coastal sediments (Jäntti et al., 2 2011). Ammonium oxidation linked to NO2 reduction or anammox (anaerobic ammonium oxidation) is another major process in sedimentary N cycle and occurs significantly in ecosystems, where denitrification is limited 2 by availability of carbon rather than NO3 (Megonigal et al., 2004). The processes and flux of N by denitrification and anammox process in the coastal sediments of Baltic Sea are schematically shown in Fig. 4.6. Increased ocean acidification inhibits ammonia oxidation coupled with increasing rates of denitrification and anammox. Predicted future trend in nitrogen cycle mediated by microorganisms are: processes like N2 fixation, denitrification and anammox increases; nitrification decreases (Hutchins and Fu, 2017)(Fig. 4.7). The net changes will be shifted

Figure 4.6 Denitrification and anammox processes in coastal sediments of Baltic Sea (Carstensen et al., 2014). Yellow arrows indicate denitrification/anammox processes; red arrows indicate processes that leads to release of ammonia; blue arrows indicate cycling of P; numbers in brackets are the estimates of nitrogen and phosphorus sources (river input) and sinks (burial, removal and recycling) in kt of N or P per year. Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 129

(A) (B)

Figure 4.7 (A) Current and (B) future predicted changes in microbially mediated nitrogen cycle in the oceans (Hutchins and Fu, 2017). toward reduced N species such as ammonium from oxidized species such as nitrate and nitrite.

4.2.3 Sulfur Sulfur compounds cycle via between the soil, oceans, atmo- sphere, and living organisms. Sulfur cycle is an important microbial medi- ated element cycle in coastal lake sediments during OM degradation and involves both oxidative and reductive processes (Jørgensen, 1990). Sulfate reduction is a major terminal electron accepting process in anoxic coastal sediments (Holmer and Storkholm, 2001; Vincent et al., 2017). Natural sources of sulfur in coastal systems are from weathering of sulfur- containing rocks and the oxidation of organic sulfur from terrestrial sources. In sediments, sulfate is supplied from overlying water and also produced during in situ hydrolysis of sulfate esters and reoxidation of reduced sulfur species by various oxidants (Schippers and Jørgensen, 2002). Nevertheless, human activities such as mining and fossil fuels burn- ing have altered the balance of this cycle, wherein inputs of sulfur com- pounds in surface waters and atmosphere have increased extensively (Friese et al., 1998). Sulfur dioxide in the atmosphere is deposited as acid rain, which also contributes to sulfate concentration, in addition to acidifi- cation (David and Mitchell, 1985). The concentration of sulfur in seawa- ter is 28 mM; in freshwater, it ranges from B10 to .500 μM; in 130 Microbial Communities in Coastal Sediments oligotrophic lakes, the concentration is ,300 μM; and in eutrophic lakes, the concentration is up to 700800 μM. Sulfur cycling is more prominent in coastal and marine systems than terrestrial and freshwater environments due to the abundance of sulfate. Sulfur reduction is a major TEA process during OM mineralization in coastal sediments due to the presence of high concentration of sulfate in sediment pore water. Enhanced sulfate input stimulates sulfate reduction as a predominant process, thus altering the cycling of other elements such as carbon, nitrogen, and iron (Cook and Kelly, 1992). A large proportion of pore-water sulfate reduction occurs by coupling to anaerobic methane oxidation pathway (Borowski et al., 1996). Sulfate-reducing bacteria (SRB) have a key role in sulfur cycle, where they reduce sulfate to H2S by utilizing sulfate as TEA under anaerobic conditions during the degradation of OM. Sulfate reduction is significant in eutrophic conditions and less significant in oligotro- phic conditions, due to higher availability of OM and sulfate content in the former. Deposition of sulfur is controlled by sulfate reduction as well as reoxidation. Reoxidation occurs under oxic conditions in the presence of sulfide-oxidizing bacteria (Fig. 4.8). Sulfur deposition is high in eutrophic than oligotrophic conditions due to less oxygen concentration in the sediment and consequently reduced reoxidation. This process is also attributed to less faunal activity in the sediments (Holmer and Storkholm, 2001). The gaseous sulfur compounds that are either produced or con- sumed by microorganisms in coastal sediments and that have an important role in the global sulfur cycle are H2S, dimethyl sulfoxide (DMSO), dimethyl sulfide (DMS), methane thiol (MeSH), carbonyl sulfide (OCS), carbon disulfide (CS2), and dimethyl disulfide (DMDS) (Bates et al., 1992; Lomans et al., 2002). Among these, the primary organosulfur gas is H2S that is produced by dissimilatory sul- fate reduction (Bremner and Steele, 1978). Several group of microor- ganisms transform sulfur compounds like DMSO to DMS and vice versa. DMS contributes to approximately 75% of the sulfur that enters the atmosphere and accounts for approximately 90% of the biogenic sulfur emission from coastal marine environments. Dimethyl sulfonio- propionate (DMSP) produced by marine micro- as well as macroalgae and halophytic plants are the source of DMS in marine systems (Kiene, 1996). Coastal marine sediments that are highly reduced and sulfide-rich due to less oxygen penetration and sulfide precipitation Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 131

Figure 4.8 Microbe-mediated sulfur cycle (Muyzer and Stams, 2008). are referred to as sulfureta-like sediments that contribute to signifi- cant release of H2Stotheatmosphere(Megonigal et al., 2004).

4.2.4 Manganese (Mn) and iron (Fe) The most dynamic and geochemically important elements on earth, Mn and Fe, are involved in a variety of oxidationreduction process in aquatic sediments. They exist in a variety of forms such as amorphous, crystalline, dissolved, adsorbed, precipitated, and reduced forms (Berner, 1981; Friedl et al., 1997). Oxidized forms of Mn(IV) and Fe(III) oxides with low solubility are found in the surface sediments; whereas, the reduced forms such as Mn(II) and Fe(II) are more soluble and found in the reduced deeper sediments (Thamdrup and Canfield, 1996; Thamdrup, 2000). In coastal sediments, effective manganese fluxes are higher than effective iron fluxes to approximately one order of magnitude. 132 Microbial Communities in Coastal Sediments

Reoxidation of manganese and iron oxides during oxygen-deprived con- ditions in the sediment leads to their precipitation, thus contributing to OM degradation (Middelburg and Levin, 2009). Utilization of manganese oxides by microbes as TEAs is thermodynamically favorable than iron oxi- des. Hence, depletion of manganese stocks occurs at higher oxygen levels than iron (Kristiansen et al., 2002). The manganese oxide pool may be depleted within a short period of few days, if the flux of dissolved manga- 1 nese is not balanced by reoxidation of Mn2 . However, depletion does not occur as the stores are refurbished from external sources like high- input metal oxiderich material eroded from adjacent areas and from riverine sources. From the overlying water column, surface Mn and Fe oxides are built up rapidly, particularly when oxic conditions are estab- lished between periods of anoxia (Slomp et al., 1997). Two groups of iron-oxidizing bacteria have major role in oxidizing iron. The first group is photosynthetic bacteria that use sunlight to fix CO2 to organic carbon under anaerobic conditions, wherein, ferrous iron is the electron donor (Ehrenreich and Widdel, 1994).

ðÞ1 1 Sunlight! ðÞ1 ðÞ Fe II CO2 H2O Fe III CH2O n The second group is chemolithotrophic bacteria, where ferrous iron is used as a source of energy and nitrate as the oxidant for anaerobic growth (Straub et al., 1996). Dissimilatory metal reduction of iron or manganese has been reported in Geobacter metallireducens, where the bacteria grow het- erotrophically by the reduction of iron or manganese oxides. A hypotheti- cal of iron, involving oxidation and reduction process to complete the iron cycle is given as Fig. 4.9 (Nealson, 1997). A major difference between sedimentary iron and manganese cycling is that the interactions with sulfur cycle are stronger for iron and weak for

Figure 4.9 Microbe-mediated iron cycle (Nealson, 1997). Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 133 manganese. Part of sulfide generated during sulfate reduction reacts with reactive iron, which may be dissolved or precipitated to form iron sulfides and S0 (Thamdrup et al., 1994), which is a common and important bio- geochemical process in sediments (Berner, 1984). These intermediate sul- fur compounds stimulate pyrite formation (Damsté and de Leeuw, 1990). The iron sulfide minerals formed in coastal sediments are oxidized effi- ciently and then integrated into the oxidative sedimentary S cycles (Jørgensen, 1977). Precipitation of manganese sulfides is unusual. Mn(IV) reduction is either a microbially mediated organotrophic respiration pro- cess or a spontaneous abiotic reoxidation process coupled to reoxidation of reduced S and Fe that act as electron donors (Kristensen, 1988).

4.3 Greenhouse gas dynamics in coastal ecosystems Coastal sediments are considered as potential sinks of atmospheric carbon. Nevertheless, improper management may cause them to act as sources of GHGs such as carbon dioxide and methane. For example, intense human activity such as agriculture and sewage disposal in coastal zones also leads to nutrient input into coastal ecosystem, which stimulates eutrophication. In addition, tropical conditions exhibit strong precipitation patterns and high temperature resulting in excessive nutrient input into coastal ecosys- tems during monsoon and increased degradation of OM in the sediments. For example, cumulative terminal electron accepting processes in a tropi- cal coastal lake were higher during summer, which reveal the influence of temperature on heterotrophic microbial activity in coastal lake sediments (Vincent et al., 2017).

4.3.1 Carbon dioxide Although coastal sediments sequester carbon, when disturbed or warmed, they release the major heat-trapping GHGs such as CO2 or CH4 (Moomaw et al., 2018). Moreover, conversion of coastal ecosystems to agriculture has been done in many countries, which allows soil OM to be oxidized and releases CO2 to the atmosphere. For example, the removal of mangroves for aquaculture or coastal development is one of the major causes. Nearly one-third of CO2 added to the atmosphere from human activity is due to deforestation and oxidation of disturbed sediment OM. Carbon dioxide is formed during aerobic and anaerobic degradation of OM. In coastal ecosystems, the OM produced in the water column (autochthonous) or leached from the catchment (allochthonous) settles 134 Microbial Communities in Coastal Sediments and is mostly remineralized in sediments; some of the settling OM is pre- dominantly degraded by aerobic microbial processes, which provide maxi- mum energy yield. The electron acceptors that diffuse into the sediment from the overlying water are consumed in sediments in the order of their energy yield capacity. This results in vertical zonation of various minerali- zation processes (Conrad, 1996). CO2 is formed during aerobic and anaerobic degradation of OM. CO2 is produced during anaerobic fermentation process by acetoclastic methanogens that utilize acetate as substrate (Liikanen et al., 2003). The occurrence and ratio of oxidized and reduced compounds regulate sedi- ment redox potentials, which can be used as an indicator of the most dominant degradation process. CO2 is produced during anaerobic fermen- tation processes and acetoclastic methanogenesis. CO2 can also be con- verted to methane during methanogenesis or to acetate during homoacetogenesis in anaerobic sediments (Schulz and Conrad, 1996). As CO2 is a water-soluble gas, it may be transported from the atmosphere into the sediments through water column and also to the atmosphere from sediments. In water column, CO2 can be fixed back to biomass in primary production. Coastal ecosystems can be either net sources or sinks of atmospheric CO2 depending on the surface water CO2 concentration in respect of atmospheric equilibrium. CO2 exchange has a strong diurnal and seasonal variation within an ecosystem, thus making them to act both as a sink or source of CO2 depending on the time and season (Anderson et al., 1999). The variation in CO2 exchange is regulated mainly by radia- tion and photosynthetic activity. Thus coastal ecosystems can act as sinks or sources of CO2, which is dependent on diurnal seasonal variation. When heterotrophy dominates autotrophy, organic carbon is degraded by sediment microorganisms and CO2 is released. Conversely, when autotrophy dominates heterotrophy, carbon is accumulated in sediments (Cole, 1999; Schindler et al., 1997) and every year nearly 2.6 6 0.5 Pg C, which is equal to 25% of the annual emissions is removed by the primary productivity of phytoplanktons and dissolution in the ocean waters (Fig. 4.10)(Moomaw et al., 2018). Interestingly, diurnal variations are also exhibited in CO2 emissions from coastal sediments. Due to primary productivity, coastal ecosystems are sinks of CO2 during daytime and source during nighttime. Moreover, CO2 can also be reduced to acetate of CH4 in methanogenesis in the sediments; whereas, in the water column, CO2 is fixed by primary pro- duction forming biomass (Lay et al., 1998). CO2 exchange between the Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 135

Figure 4.10 Global carbon dioxide budget (Moomaw et al., 2018). wetlands and atmosphere varies spatially and temporally, showing strong diurnal and seasonal variations.

4.3.2 Methane Methanogenesis is the final step in the anaerobic degradation of OM. Most of the OM (up to 87%) is degraded by various fermentation pro- cesses as well as by methanogenesis to methane and carbon dioxide in anoxic sediments (Capone and Kiene, 1988). The primary steps in metha- nogenesis are conversion of acetate to CO2 and CH4.CH4 is produced biologically by methanogenesis, which is the pathway for mineralizing organic carbon compounds in anoxic conditions and substrate availability (Laanbroek, 2010). Methanogenic microorganisms, primarily Archaea or Methanogenic Archaea, occur in anaerobic environments such as coastal sediments (Schulz and Conrad, 1996) and play an important role in OM 136 Microbial Communities in Coastal Sediments

degradation leading to CH4 production. Methanogens utilize only a lim- ited number of small organic molecules in energy production. Methane can be produced by reduction of CO2 to CH4, in which H2, formate, alcohol, or CO are used as electron donors, that is, as reductants (Boone, 1991). Contrastingly, in freshwater sediments, most of the methane is pro- duced by the reduction of methyl group of acetate (Woltemate et al., 1984; Whiticar et al., 1986; Jones, 1991). Methanogens also reduce the methyl group from methanol, dimethyl sulfide, mono-, di-, and trimethy- lamines (Jones, 1991). These methylated “noncompetitive substrates” are used especially in marine sediments, compete with methanogens for H2 and other low-molecular-weight substrates, such as acetate and formate (Capone and Kiene, 1988; Kiene, 1991). Although coastal wetlands are potential zones for CH4 production, the overall environmental and ecological factors influencing the rate of methane production vary temporally and spatially (Purvaja and Ramesh, 2001; Gonsalvesetal.,2011). Methanogenesis is influenced by several environ- mental factors such as salinity, temperature, pH, microbial interactions (Marinho et al., 2012), availability of sulfate, and redox potential (Grünfeld and Brix, 1999). In tropical conditions, increased input of nutrients and OM due to monsoon runoff triggers methanogenic activity (Reshmi et al., 2015). In coastal sediments, methane emissions are reduced due to competition for common substrates by SRB, in which case, SRB outcompete methanogens. In addition, the presence of aquatic macrophytes significantly influences CH4 dynamics (Whiting and Chanton, 1993; Laanbroek, 2010). With increase in trophic status, accumulation of OM and nutrients also increases (Deevey et al., 1986). Hence, emission of methane is usually high from eutrophic ecosystems with anoxic hypolimnions, whereas the methane emis- sions from oligotrophic systems with a good O2 status can be negligible due to low methane production and effective methane oxidation in surface sedi- ments and water column. Methane is transported to coastal ecosystems by diffusion, advection, ebullition, or through plants (Chanton and Whiting, 1995). Aquatic macrophytes are important in conducting methane in their aerenchyma directly from the anoxic sediments into the atmosphere (Dacey and Klug, 1979). Plants also provide fresh, easily decomposable OM in root exudates into the sediments, thus promoting methanogenesis. At the same time, plants provide oxygen for root respiration that increases CH4 oxidation in their vicinity (Chanton and Whiting, 1995). CH4 production is an important environmental issue in the context of concerns about global climate change. The relative contributions of Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 137

different natural sources to global atmospheric CH4 emissions are 76% from wetlands, 11% from termites, 8% from oceans, and 5% from hydrates (IPCC, 2001). Moreover, the relative increase of CH4 in the atmosphere since preindustrial period is approximately 150%, compared to 35% increase for CO2.CH4 emissions from wetlands are between 177 and 21 284 Tg CH4 year .CH4 has a 100-year global warming potential more than 28 times that of CO2 (Myhre et al., 2013). In the water column, CH4 is oxidized aerobically by methanotrophs that utilize methane as a carbon source. Methanotrophs use CH4 as electron donor and carbon source. CH4 is first oxidized to methanol and further to CO2. Anaerobic oxidation of methane is coupled to sulfate reduction or iron(III) or manga- nese(IV) reduction. The global methane budget (Fig. 4.11) reveals signifi- cantly lesser contribution by coastal ecosystems (Moomaw et al., 2018).

Figure 4.11 Global methane budget (Moomaw et al., 2018). Black arrows indicate natural emissions; red arrows indicate anthropogenic emissions since 1750; brown arrows indicate both natural (natural fires) and anthropogenic (biomass burning). 138 Microbial Communities in Coastal Sediments

4.3.3 Nitrous oxide

Nitrous oxide is mainly produced in denitrification, where N2O is pro- duced as an intermediate in the reduction of NO3 to N2 (Firestone and Davidson, 1989). Denitrification is an anaerobic microbial respiratory 2 2 pathway based on nitrogen, where nitrogen oxides (NO3 or NO2 ) are reduced to nitrogen gases such as N2OorN2. During denitrification, energy is conserved by the microbial cell by coupling electron transport 2 2 phosphorylation to the reduction of NO3 or NO2 . During this pro- cess, nitrogen is not assimilated into the cell and hence termed as dissimi- latory nitrate reduction. This process is thermodynamically favorable and predominates other respiratory metabolic processes such as Fe(III) reduc- tion and sulfate reduction or methanogenesis when nitrogen oxides are available in the sediments. Sediments serve as important sites for denitrification (Chan and Campbell, 1980), nitrification (Hall, 1986), and N2O production (Mengis et al., 1996). However, nitrification is generally limited by the availability of oxygen. Nitrous oxide production is measured by acetylene inhibition method and 15Nor13N tracer method and the latter is used in recent studies. Denitrification can be uncoupled from nitrification when there is sufficient external supply of NO3. Meanwhile, uncoupled nitrifica- 1 tiondenitrification occurs in the sediment in the presence of NH4 and 2 organic carbon, with limited external input of NO3 . Anoxic sediments with an abundant supply of organic carbon provide a good environment for heterotrophic denitrification. N2O production may increase as a result of eutrophication or acidification due to the inhibition of N2O reduction by NO3 and H2S(Seitzinger, 1988). In the water column, nitrification produces N2O in the oxic surface waters (Elkins et al., 1978; Prisu et al.,

Figure 4.12 Microbe-mediated pathways of nitrification and denitrification (Megonigal et al., 2004). Biogeocycling of nutrients (C, N, P, S, and Fe) and implications 139

Figure 4.13 Biological interactions and nitrous oxide production (Stock et al., 2014). (1) The excretion products of copepods increase the carbon availability for DNRA 2- bacteria; (2a) the increased carbon availability also enhances SO4 reduction leading to the formation of H2S; (2b) increase in H2S inhibits denitrification. (3) Diatoms influ- ence denitrification indirectly by enhancing the survival of copepods; (4) diatoms influence the quality and quantity of excretion products by copepods.

1996) and denitrification can either produce or consume N2O depending on the oxygen availability. Small-scale biological interactions between bacteria, copepods, and diatoms can have an important impact on denitri- fication and also on sediment nitrogen fluxes (Stock et al., 2014) (Figs. 4.12 and 4.13). Similar to CH4, coastal sediments are also known to 2 be either sinks or sources of N2O depending mainly on O2 and NO3 availability and the level of eutrophication. Generally, pristine ecosystems have low N2O fluxes (Mengis et al., 1997). However, the anthropogenic inputs of nitrogen significantly increase N2O emissions (Kaplan et al., 1978; Seitzinger and Kroeze, 1998) from coastal sediments.

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5.1 Introduction Coastal sediments are repositories of physical and biological debris and act as sinks for a wide variety of organic and inorganic pollutants. Chemical contaminants present in the may be immobilized and accumulated in sediments or may be subject to transformation and activa- tion processes. Chlorinated organic compounds comprise a large group of compounds with diverse toxic effects on biological systems. Studies per- taining to their biodegradation and bioremediation are scanty owing to their complex characteristics, particularly their recalcitrant nature. Chlorinated organic compounds are highly resistant to degradation and are highly lipophilic. Dioxins, dichloro-diphenyl-trichloroethane (DDT), polychlorinated biphenyls (PCBs), and some organochlorine pesticides are compounds of scientific and public concern because of their widespread occurrence in the environment and high toxicity. As a class of chemicals, organochlorines, because of the strong molecular bond made when chlo- rine combines with carbon, are difficult to break down either in the envi- ronment or living tissue. Many organochlorines are very stable, and they persist for long periods of time. Breakdown products often can be more toxic and persistent than the original product. Many pollutants such as polychlorinated aromatic hydrocarbons and polycyclic aromatic hydrocarbons (PAHs)arehighlyapolar,havelowsolu- bility in water and high tendency for strong adsorption onto organic matter in soil and sediment matrices (Means et al., 1980). Particularly, they can become dangerous if they enter the food chain, since several of the more persistent compounds, such as PAHs and PCBs are carcinogenic. When organochlorine compounds are biodegraded in anaerobic systems, biodegra- dation occurs mainly as a result of the first ring being broken or the halogen

Microbial Communities in Coastal Sediments © 2021 Elsevier Inc. DOI: https://doi.org/10.1016/B978-0-12-815165-5.00005-4 All rights reserved. 147 148 Microbial Communities in Coastal Sediments being changed by hydroxyl radicals. The resistance to biodegradation by anaerobic process increases with increase in chlorination. However, under anaerobic conditions, chlorine can be removed from the aromatic ring by reductive dechlorination, resulting in partially or fully dehalogenated pro- ducts, which are then more susceptible to either aerobic or anaerobic attack. In the past few years, several studies have shown that halogenated phe- nols are reductively dehalogenated in sewage sludge, aquatic sediments, and soils. Some of these studies indicate that reductive dehalogenation reactions may thrive in methanogenic environments. One of the difficul- ties involved in biodegrading organochlorine compounds is their toxicity to microbes that hinders the degradation process in ways that they are not yet fully understood. Nevertheless, many organochlorines have been shown to support energy-yielding metabolisms benefitting anaerobic bac- teria as the primary energy and carbon source or as electron acceptors of respiration. The extent of pollution by persistent organic pollutants (POPs) in coastal sediments and the role of bacteria particularly anaerobic bacteria involved in the degradation of these recalcitrant compounds in anoxic sediments are discussed in this chapter.

5.2 Why persistent organic pollutants? POPs can cause serious problems in various environmental compartments due to their toxicity, persistence, and . Most of them with hydrophobic properties, include simple aromatic compounds such as benzene, toluene, ethylbenzene, and xylenes (BTEX), PAHs, including naphthalene, phenanthrene, and benzopyrene, and PCBs (Semple et al., 2003; Diez, 2010). The anthropogenic sources of some persistent organic compounds are summarized in Table 5.1. The priority list of POPs has been updated recently to include com- pounds like brominated flame retardants such as polybrominated diphenyl ethers and hexabromocyclododecanes (Zacharia, 2019). Organic pollutants in sediments are a worldwide problem because sediments act as sinks for hydrophobic, recalcitrant, and hazardous compounds. Depending on bio- geochemical processes, these hydrocarbons are involved in absorption, desorption, and transformation processes and can be made available to benthic organisms as well as organisms in the water column through the sedimentwater interface. Most of these recalcitrant hydrocarbons are toxic and carcinogenic; they may enter the food chain and accumulate in biological tissue (Perelo, 2010). Biodegradation and biotransformation of persistent organic pollutants 149

Table 5.1 Sources of persistent organic compounds (Field et al., 1995). Organic compounds Industrial source BTEX Fossil fuels, solvents, industrial feedstocks Styrene Plastics PAH Fossil fuels, wood preservation Alkyl phenols Surfactants, detergents Aromatic sulfonates Surfactants, detergents, sulfite pulping, dyes Aromatic amines Pesticides, dyes, pigments, pharmaceuticals Azo aromatics Dyes Nitroaromatics Explosives, pharmaceuticals, pesticides, dyes Chlorophenols and dioxins Wood preservation, pesticides, pulp bleaching effluents Chloroaromatic hydrocarbons Pesticides, solvents, dielectric and hydraulic and PCB fluids

BTEX, Benzene, toluene, ethylbenzene, and xylene; PAH, polycyclic aromatic hydrocarbons; PCB, polychlorinated biphenyls.

In the sediments, organic pollutants may associate temporarily in the particulate matter, establishing equilibrium relations in the watersediment interface. These sorption and desorption processes substantially determine the bioavailability of toxic substances. The direct transfer of chemicals from sediments to organisms is now considered to be a major route of exposure for many species. To evaluate the contaminant release from sediment through desorption processes, both the characteristics of the sediment and the overlying water column must be considered (Zoumis et al., 2001). Contaminants may be mobilized by changes in geochemical parameters, by diffusion of pollutants into the water body because of concentration gradi- ents, by oxidation of anoxic sediments through bioturbation or resuspension caused by flooding, as well as by degradation processes leading to a more mobile form.

5.3 Anaerobic degradation and pathways Biodegradation is a multistep process that is taking place in the presence of a number of microorganisms that often act synergistically. The biodeg- radation rate in real coastal sediments depends on physicochemical charac- teristics of the sediments such as the presence of particulate matter, concentration of inorganic and organic nutrients, temperature, oxygen concentration, redox potential as well as the properties and “age” of the pollutant. Moreover, the range and rate of biodegradation processes 150 Microbial Communities in Coastal Sediments depend on several factors such as the adaptation of native microbial popu- lation, composition, and activity of bacterial flora present in the medium in which the reaction occurs (Gotvajn and Zagorc-Koncan,ˇ 1999). In general, organic compounds are susceptible to biodegradation both under aerobic and anaerobic conditions. Since anaerobic degradation mechanisms do not involve molecular oxygen, aromatic compounds with functional groups (benzene, toluene, naphthalene, etc.) are difficult to degrade. However, higher chlorinated aromatic compounds are degraded better under anaerobic conditions. Particularly, electrophilic aromatic pol- lutants with multiple chloro, nitro, and azo groups have proven to be per- sistent to biodegradation by aerobic bacteria. These compounds are readily reduced by anaerobic consortia to lower chlorinated aromatics or aromatic amines, but are not mineralized further. The reduction increases the susceptibility of the aromatic molecule for oxygenolytic attack. Microorganisms in comparison to other organisms have a particular predisposition to adapt to novel environmental conditions and the ability to utilize compounds that are not the products of their own metabolism, as substrates needed for energy production and structure building. One molecule of enzyme can catalyze decomposition of millions of organic molecules per minute. The reactions mediated by microorganisms are to a large degree, similar to those occurring in higher organisms. Therefore the aromatic compounds undergo epoxidation and hydroxylation, the ali- phatic ones are oxidized and degraded through β-oxidation pathway, and the nitro-organic derivatives are metabolized with the use of nitroreduc- tases. Microorganisms can also mediate the processes that the higher organisms are not capable of decomposition or aromatic ring or dehalo- genation (Dabrowska et al., 2004). Evidence suggests that the biodegradation of anthropogenic haloge- nated organic compounds has basis in two phenomena: (1) the fortuitous degradation of structures analogous or similar to naturally occurring com- pounds, using existing pathways for naturally occurring organohalides or (2) a selective genetic transfer, amalgamation, or mutation, so that new pathways evolve from existing ones. Microorganisms have evolved a vari- ety of metabolic strategies for cleaving the carbonhalogen bond. These biodegradation mechanisms include oxidation, reduction, substitution, intramolecular substitution, dehydrohalogenation, hydration, and methyl transfer reactions (Häggblom and Bossert, 2004). Reductive dehalogenation is considered to be the predominant process in the anaerobic transformation of halogenated compounds. In addition to Biodegradation and biotransformation of persistent organic pollutants 151 potentially serving as a carbon source, organohalides function as terminal electron acceptors in an anaerobic respiration process, termed dehalorespira- tion or halorespiration (Häggblom and Bossert, 2004). Dehalogenation reactions comprise different strategies, where organohalides serve either as electron donors (and carbon sources) or electron acceptors, or undergo exchange reactions as follows: 1. the organohalide serves as a carbon and energy source and dehalogena- tion occurs in order to break down the carbon backbone, 2. the organohalide serves as an alternate electron acceptor for anaerobic respiration termed (de)halorespiration, 3. dehalogenation occurs as a detoxification mechanism, or 4. the organohalide is dehalogenated through fortuitous reactions that do not yield any benefit to the organism. Anaerobic transformation of chlorinated organic compounds involves reductive dehalogenation where the halogenated organic compound serves as the electron acceptor and the halogen constituent is replaced with hydrogen (Morris et al., 1992). R-Cl 1 2e2 1 H1-R-H 1 Cl2

5.3.1 Phenols and chlorinated phenols Anaerobic phenol degradation has been elucidated in detail with denitrify- ing bacteria (Fuchs et al., 1994). Phenol is first carboxylated to 4- hydroxybenzoate, a reaction in which phenylphosphate is an intermediate (Lack and Fuchs, 1994). Subsequently, 4-hydroxybenzoate is activated to 4-hydroxybenzoyl-CoA, which is reductively dehydroxylated to benzoyl- CoA. The aromatic ring is reduced (Koch and Fuchs, 1992), and subse- quently the ring is further metabolized by β-oxidation (Fig. 5.1).

Figure 5.1 Pathways involved in the anaerobic degradation of phenol (Fuchs et al., 1994). 152 Microbial Communities in Coastal Sediments

Miller et al. (2004, 2008) explored the 40-year history of Trichlorosan (TCS) and Triclocarban (TCC) deposition in estuarine sediment at two locations on the US East Coast. Data showed that (1) TCC and to a lesser extent TCS are persistent organic contaminants of estuarine sediments; (2) TCC is more persistent and more abundant than TCS; (3) in aged sed- iment, TCC can undergo slow anaerobic dechlorination and the process shows geographic variability; (4) anaerobic transformation processes can alter the chlorine substitution pattern but do not reduce the overall quan- tity of carbanilide species present; and (6) TCC contamination of estuarine sediment in some locations extends into the ppm range, representing potentially unhealthy levels for aquatic organisms. Further studies into the aquatic toxicity of these persistent compounds are needed to more accu- rately judge their actual threat to aquatic ecosystems.

5.3.2 3-Chlorobenzoate In the anaerobic transformation of 3-chlorobenzoate under methanogenic conditions, a consortium of bacteria is involved (Mohn and Tiedje, 1992). A dechlorinating bacterium, Desulfomonile tiedjei, was isolated from 3- chlorobenzoate-degrading methanogenic consortia, which was able to gain energy for growth by reductive dechlorination. The reducing equiva- lents needed for this reduction are obtained through interspecies H2 trans- fer originating from another bacterium in the consortium that oxidizes benzoate to acetate, CO2, and H2. Fig. 5.2 shows the pathways involved in the degradation of 3-chlorobenzoate under methanogenic conditions. The benzoate-degrading bacterium is only able to catabolize benzoate when all the reducing equivalents (hydrogen and/or formate) are con- sumed by the methanogens, together with the dechlorinating bacterium. A novel alphaproteobacterium affiliated with the genus Magnetospirillum

Figure 5.2 Pathways involved in the degradation of 3-chlorobenzoate under metha- nogenic conditions (Mohn and Tiedje, 1992). Biodegradation and biotransformation of persistent organic pollutants 153

isolated from sediments was able to degrade 4-methylbenzoate to CO2 under nitrate-reducing conditions (Lahme et al., 2012).

5.3.3 Polycyclic aromatic hydrocarbons PAHs are ubiquitous pollutants. There are over 100 different PAH com- pounds. PAHs are mainly formed by incomplete combustion of organic compounds and rarely are of industrial use, except for a few PAHs used in medicines and the production of dyes, plastics, and pesticides. They are highly toxic to organisms due to their carcinogenic and mutagenic poten- tial. Because of their low water solubility and hydrophobicity, they tend to adsorb on and accumulate in sediments, where the degradation of PAHs with high molecular weights is particularly slow (Readman et al., 1982, 2002). During the past few decades, several studies have revealed the role of microorganisms in deciding the fate of hydrocarbon pollution in sediments. Benthic ecosystems act as a repository for oil contamination and exhibit pronounced responses from microbial abundance, diversity as well as community composition in the sediments. Nevertheless, benthic microbial communities were resilient to such sedimentary changes and recovered quickly after oil disturbance, thus maintaining ecosystem func- tions (Kostka et al., 2011). Moreover, the microbial degradation of hydro- carbon in coastal sediments also depends on physicalchemical factors such as electron acceptors, nutrients and also sediment biology such as sediment reworking and bioturbation activity (Fig. 5.3). Rapid and complete degradation of PAH occurs under aerobic condi- tions. The initial degradation process is oxidative catalyzed by oxygenases and peroxidases (Fig. 5.4). Further peripheral degradation pathways con- vert the organic compounds to intermediates in a stepwise process. The intermediates then enter into the central intermediary metabolism of the bacteria such as the tricarboxylic acid cycle (Das and Chandran, 2011). The degradation of PAH by aerobic microorganisms has been extensively studied; however, many natural environments contaminated with PAH are anoxic, for example, the coastal sediments. In these habitats, anaerobic degradation plays an important role in the health of the biosphere. Studies pertaining to anaerobic degradation of PAH are less and relatively recent, wherein the degradation of PAHs such as naphthalene, anthracene, phen- anthrene, fluorine, acenaphthene, and fluoranthene has been reported in anaerobic conditions (Mallick et al., 2011). 154 Microbial Communities in Coastal Sediments

Figure 5.3 Fate of PAH in coastal sediments.

Figure 5.4 Aerobic biodegradation of PAH. Biodegradation and biotransformation of persistent organic pollutants 155

Anaerobic microorganisms are able to reductively dechlorinate poly- chlorinated aromatic hydrocarbons (Häggblom, 1992). The highly elec- trophilic character of the multiple chlorine substitutions increases the favorableness of this type of nucleophilic attack (Dolfing and Harrison, 1993). Anaerobic enrichment cultures were able to dechlorinate hexa- and pentachlorobenzene two to three times faster than tetra- and trichlor- obenzenes (Holliger et al., 1992). Reductive dechlorination can also be carried out abiotically with reduced metal cofactors (e.g., vitamin B12 and factor F430) of important enzymes involved in anaerobic metabolism. Reductive dehalogenation of polychlorinated aromatic hydrocarbons does not result in the mineralization of these compounds. The anaerobically recalcitrant end products of this biotransformation include lower chlori- nated analogues. Contrastingly, the di- and monochlorinated phenols, which often accumulate from the reductive dehalogenation of polychlori- nated phenols, undergo further metabolism to mineralized products. Verification of anaerobic degradation of PAHs under nitrate-reducing conditions has been presented by Bregnard et al. (1996) and Langenhoff et al. (1996). Chang et al. (2001b) suggest that anaerobic microorganisms might have greater potential for organic-pollutant detoxification in the environment. Sulfate-reducing bacteria constitute a large fraction of bacte- ria in oil-contaminated sediments and sulfate reduction was the predomi- nant electron accepting process (Andrade et al., 2012; Cravo-Laureau and Duran, 2014). Comparison of phenanthrene degradation under three reducing conditions was done and it has been the order of phenanthrene remaining for sediment sample: nitrate-reducing conditions . sulfate- reducing conditions . methanogenic conditions. It has also been found that additional acenaphthene and phenanthrene were completely degraded in sediments within a 56-day incubation, while pyrene, fluorine, and anthracene degraded only 4.0%, 28.0%, and 48.7% within a 56-day incu- bation period. The schematic pathway of anaerobic degradation of aro- matic compounds is given in Fig. 5.5. Diverse aromatic compounds are first transformed into central intermediates (benzoyl-CoA, resorcinol, or phloroglucinol) that are subsequently reduced to alicyclic compounds. The ring is then cleaved by hydrolysis and the noncyclic products are transformed into the central metabolite acetyl-CoA by β-oxidation. As a result of the hydrophobic nature of PAHs, sediments are the primary repository of these compounds, particularly in the marine envi- ronment (Harris et al., 2011). Novel PAH dioxygenase gene variants were present in abundances similar to or higher than those of phnA1 from 156 Microbial Communities in Coastal Sediments

Figure 5.5 Schematic pathway of anaerobic degradation of aromatic compounds (Fuchs et al., 1994).

Cycloclasticus sp. at a chronically polluted subantarctic coastal marine environment in Patagonia. These novel gene variants were detected over a 6-year time span and were also present in sediments from temperate Patagonian sites (Marcos et al., 2012).

5.3.4 Polychlorinated biphenyls PCBs are among the worst pollutants because of their toxicity, carcinoge- nicity, wide distribution, and slow biodegradation in the environment (Meagher, 2000). PCBs are used in hydraulic fluids, plasticizers, adhesives and lubricants, flame retardants, and dielectric fluids in transformers. They are released during production from spillage and disposal (Scragg, 2005). It is reported that thousands of metric tons of commercial PCBs persist in aquatic sediments. Biodegradation of PCBs is a multistep process that involves aerobic and anaerobic bacteria. Anaerobic bacteria are capable of decomposing compounds containing several chlorine atoms while aerobic bacteria only degrade compounds with one or two chlorine atoms. Anaerobic dechlorination has been observed in a large number of sedi- ments (Abramowicz, 1990; Brown et al., 1987). The rate, extent, and pattern of dechlorination of four Aroclors by inocula prepared from two PCB-contaminated sediments were compared (Quensen et al., 1990). The four mixtures used, Aroclors 1242, 1248, 1254, and 1260 average approximately three, four, five, and six chlorines, respectively, per biphenyl molecule. All four Aroclors were dechlorinated with the loss of meta plus para chlorines ranging from 15% to 85%. It was previously noted that greater dechlorinating activity was associated with Biodegradation and biotransformation of persistent organic pollutants 157

PCB-contaminated sediments. This implies that there has been selection of contaminated sites for microorganisms capable of effecting dechlorina- tion. The responsible selective pressures may arise from the ability to use PCBs as terminal electron acceptors and/or from the ability to use the energy that is potentially available from dechlorination (Quensen et al., 1990). Beurskens et al. (1993) have estimated a half-life time of 9 years for PCB in the anaerobic Ketelmeer sediment. Reductive dechlorination primarily takes place in the meta- and para-positions and as a consequence, there is a tendency for the accumulation of ortho-substituted PCBs proces- sing one, two, or three chlorine groups (Alder et al., 1993). PCBs with as few as three chlorine groups can be dechlorinated in anaerobic sediments (Boyle et al., 1993). In complex mixtures of PCBs, most mono- and dichlorinated congeners are recalcitrant to anaerobic biodegradation. There are several reports of anaerobic dechlorination of highly chlorinated dioxins, for example, heptachlorinated dioxin is dechlorinated to hexa- and pentachlorodibenzo(p)dioxins. Methanogenic enrichment cultures acclimatized to the dechlorination of hexachlorobenzene were able to convert 1,2,3,4-tetrachlorodibenzo(p)dioxin to tri-, di- and even mono- chlorinated dioxins (Toussaint et al., 1992). Williams (1994) reported the dechlorination sequence in six trichloro- biphenyls with all chlorines in one ring. Dechlorination of every trichlor- obiphenyl occurred in sediments with all meta- and para-chlorines removed but no ortho-dechlorination was observed. The enrichment of microorganisms from sediments with 2,3,4,5,6-pentachlorobiphenyls resulted in a sequential meta- and para-dechlorination of Aroclor 1260 (Bedard et al., 1996). Two different enrichment cultures (BK 24 and BK 27) previously enriched from marine sediments with a history of PCB contamination was able to dechlorinate four octachlorobiphenyls and three nonachlorobiphenyls. The predominant dechlorination pattern showed meta-dechlorination of singly flanked m-chlorines. Some ortho- and para-dechlorination were also observed (Kuipers et al., 1999). Dechlorination rates for three primary congeners under different reducing conditions occurred in the following order (from the fastest to the slowest): methanogenic condition . sulfate-reducing condition . ni- trate-reducing condition (Chang et al., 2001a). Under methanogenic and sulfate-reducing conditions, dechlorination rates were enhanced by the addition of lactate, pyruvate, or acetate, but decreased as a result of the addition of manganese oxide or ferric chloride. Under nitrate-reducing conditions, dechlorination rates decreased by the addition of lactate, 158 Microbial Communities in Coastal Sediments pyruvate, acetate, manganese oxide, or ferric chloride. The dechlorination of the three PCB congeners is affected by changes in pH, temperature, and the presence of an electron donor or acceptor. Electron acceptors are generally the factors limiting metabolism in anaerobic environments. Thus any microorganism that could use PCBs as terminal electron acceptor would be at a selective advantage (Brown et al., 1987). Under anaerobic condition, reductive dechlorination of PCBs occurs in soils and sediments. Different microorganisms with distinct dehalogenating enzymes are responsible for different dechlorination activities and dehalogenation routes. The rate, extent, and route of dechlorination are dependent on the composition of the active microbial community, which in turn are influenced by environmental factors such as availability of carbon sources, hydrogen or other electron donors, the presence or absence of electron acceptors other than PCBs, temperature, and pH (Wiegel and Wu, 2000). The use of organic substrate as electron donors showed to increase the rate of dechlorination in Aroclor (Nies and Vogel, 1990). Anaerobic PCB dechlorination reduces the potential risk and potential exposure to PCBs in coastal sediments. In situ bioaugmenta- tion studies on contaminated sediment showed a decrease in 80% of PCBs compared to non-bioaugmented sediments, where intrinsic bioremedia- tion showed 25% decrease in PCB contamination (Payne et al., 2013). Carcinogenic potential of PCBs correlates with total chlorine levels. The decrease in risk is manifested in two ways. First, lightly chlorinated congeners produced by dechlorination can be readily degraded by indige- nous bacteria. Second, dechlorination significantly reduces bioconcentra- tion potential of the PCB mixture through conversion to congeners that do not significantly bioaccumulate in the food chain. Microorganisms are also reported to utilize PCB as electron acceptor that enables subsequent degradation (Edwards and Kjellerup, 2013).

5.3.5 Polychlorinated dibenzo-p-dioxins and dibenzofurans Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) have been deposited decades ago and are still found in deep sediment layers. PCDD/Fs are often considered recalcitrant toward biotic and abiotic deg- radation processes (Uchimiya and Masunaga, 2007). Thus they are among the most notorious environmental pollutants. Some congeners, particu- larly those with lateral chlorine substitutions at positions 2, 3, 7, and 8 are extremely toxic and carcinogenic in humans (Kaiser, 2000). Biodegradation and biotransformation of persistent organic pollutants 159

First-order half-life times of anaerobic biodegradation of some PCDD and PCDF congeners have been obtained by field experiments (Segstro et al., 1995). Contaminated sediments and anaerobic microorganisms enriched from different sediments have been used in experiments to study the persistence of PCDDs and PCDFs in sediment (Beurskens et al., 1995; Adriaens and Grbic’-Galic, 1994). The half-life times for PCDDs are gen- erally more than 100 years (Kjeller and Rappe, 1995), calculated based on sediment core measurements from the Baltic Proper. Biostimulation with electron donors, cosubstrate, and/or bioaugmentation with Dehalococcoides mccartyi was shown to enhance bioremediation of dioxins in sediments (Liu et al., 2013).

5.4 Anaerobic microorganisms involved Coastal sediments, because of their inherently anaerobic conditions and abundant carbon and energy sources, tend to have ample microbial bio- mass and diversity, which are potentially able to degrade organic pollu- tants (Himmelheber et al., 2007). A higher diversity of microorganisms has been observed in the microbial biocenoces associated with sediments than in groundwaters originating in formations of large pore size and inhabited by mobile microorganisms. Microorganisms residing in sedi- ments are sedentary, permanently bound, and living in pores of small diameter. The persistence and potential impact of some compounds in natural and engineered environments may be controlled to a greater extent by the adaptation period than by the rate of biodegradation follow- ing adaptation. Despite its importance, current understanding of the events that lead to microbial community adaptation to pollutants is extremely limited. This lack of understanding means that we are unable to predict when or where adaptive events will occur (Becker et al., 2006). The microbial species involved in biodegradation may belong to both aer- obic and anaerobic groups (Megharaj et al., 2014); however, compared to aerobic, anaerobic process is more energy saving. Intrinsic microbial con- sortia constituting dehalogenating bacterial communities have an impor- tant role in degradation of chlorinated compounds in sediments (Kuokka et al., 2014). Anaerobic dechlorination can attack a large array of chlorinated ali- phatic and aromatic hydrocarbons. During anaerobic dechlorination or reductive dechlorination, the perchlorinated substances act as terminal electron acceptors by the process of dehalorespiration (Liu et al., 2013; 160 Microbial Communities in Coastal Sediments

Zhen et al., 2014). Several anaerobic dechlorinating bacteria have been isolated (Holliger et al., 1998). These include Desulfomonile tiedjei, Desulfitobacterium, Dehalobacter restrictus, Dehalospirillum multivorans, Desulforomonas chloroethenica, D. mccartyi, Dehalococcoides ethenogenes, and the facultative anaerobes Enterobacter strain MS-1 and Enterobacter agglomerans. Some of these microorganisms reductively dechlorinate the chlorinated compound in a cometabolism reaction; others utilize the chlorinated com- pounds as electron acceptors in their energy metabolism. The characteris- tics common to dehalogenators are: (1) aryl reductive dehalogenation is catalyzed by inducible enzymes, (2) these enzymes exhibit distinct sub- strate specificity, and (3) aryl dehalogenators derive metabolic energy from reductive dehalogenation. Microorganisms with distinct dehalogenating enzymes each exhibit a unique pattern of congener activity. Some strictly anaerobic bacteria, such as Desulfitobacterium frappieri, that dechlorinate a range of aromatic chlorobenzoates or phenols also use other electron acceptors such as sulfite, thiosulfate, or nitrate when the carbon and energy sources are pyruvate (Bouchard et al., 1996). A variety of anaerobic bacteria also use halogenated compounds as carbon source for growth, in a wide range of electron accepting environments. For example, the homoacetogenic bacteria, Acetobacterium dehalogenans and Dehalobacterium formicoaceticum, can utilize chloromethane and dichloro- methane as carbon and energy sources (Mägli et al., 1996). Several chloro-, bromo-, and fluorobenzoate-utilizing denitrifying strains, repre- sentative of different groups within the Proteobacteria (e.g., Acidovorax, Azoarcus, Bradyrhizobium, Ensifer, Mesorhizobium, Ochrobactrum, Paracoccus, Pseudomonas, and Thauera) have been isolated. Their presence in many soils and sediments indicates a wide distribution of dehalogenation pro- cesses under denitrifying conditions (Song et al., 2000, 2001). In another group of bacteria, phototrophic Rhodospirillum and Rhodopseudomonas species have been shown to grow on halogenated aliphatic acids and on 3-chlorobenzoate (Kamal and Wyndham, 1990). Bedard (2008) reported dechlorination of several PCB congeners by a novel clade of putative dechlorinating bacteria within the phylum Chloroflexi. However, most of the knowledge on biodegradation of POPs by microorganisms is limited to laboratory experiments where microbial cultures were grown in con- trolled conditions (Akbari et al., 2016). The lifestyle and nutrient require- ments of these microorganisms may vary in the natural conditions depending on the prevalent environmental conditions, which suggest that more research is necessary in the field. Biodegradation and biotransformation of persistent organic pollutants 161

5.5 Limitations for anaerobic degradation: electron acceptors Depending on the prevailing redox environment and availability of elec- tron acceptors and donors, halogenated compounds offer a wide variety of substrate types that may either accept or donate electrons to microbially mediated biochemical reactions. In the absence of oxygen, different com- 1 pounds such as nitrate, sulfate, sulfur, oxidized metal ions (e.g., Fe3 and 11 Mn ), protons, and bicarbonate serve as alternative terminal acceptors. Proton and bicarbonate are the primary electron acceptors used in metha- nogenic conditions (Schink, 1992). The anaerobic degradation of organic pollutants is an important pathway in nature because many microorganisms can degrade a variety of organic compounds under anoxic conditions. However, some organic pollutants persist in waterlogged soil or sediment because of a lack of suitable electron acceptors (Wu and Marshall, 2001). For example, benzene and other aro- matic hydrocarbons are pollutants of particular concern because they are dif- ficult for microorganisms to degrade under anaerobic conditions, such as in waterlogged sediments, or groundwater. However, the addition of electron acceptors such as Fe(III) oxides, nitrates, or sulfates into contaminated sedi- ments can stimulate the anaerobic oxidation of these compounds.

5.6 Future prospects Public concern about environmental contamination, particularly by halo- genated industrial products, has led to extensive investigations of the fates of such compounds in anaerobic environments such as coastal sediments. Most of the POPs are extremely resistant to biodegradation by native microflora. The dissipation of organic recalcitrant pollutants in coastal sediments is influenced by both adsorption and biological processes. Depending on the properties, the biological decomposition of organic compounds is the most important and effective way to remove these compounds from sediments. Many studies have shown that reductive dehalogenation of aromatic as well as aliphatic compounds occurs fairly rapidly in anaerobic environments and a large number of pure cultures of anaerobes that carry out reductive dehalogenations have been isolated. At present, there is a better understanding of the microbial world’s anaerobic degradation potential owing to the studies carried out in the past few decades. Starting with initial observations of organic compound 162 Microbial Communities in Coastal Sediments disappearance in anoxic sediments, metabolically diverse microbes have steadily brought into captivity as pure cultures. The number of microbes that are present in pure cultures is nevertheless still relatively small, and the full range of microorganisms that participate in anaerobic decomposition of aromatic compounds is not entirely known. In particular, defined groups of microorganisms that normally participate in a complex web of metabolite transfers between different species have only occasionally been cultivated. Till date, attention has been diverted to biodegradation of compounds that are of interest to humans. However, it may be useful practically to try to understand biodegradation from the point of view of the participat- ing microbes. How does the possession and utilization of selected aromatic degradation pathways contribute to the success of that organism in the physical and biological environment in which it lives? With the possible exception of some halorespirers, chlorinated organic compounds are among the less favored substrates for anaerobic bacteria. Degradation enzymes are induced only under appropriate conditions of absence of oxygen and presence of substrates, and the growth rate supported is almost always substantially lower than in the presence of more favored carbon and energy source. We also need sensitive approaches for coaxing microbes to tell us more about how they perceive their environment. With the availability of complete genome sequences, differential displays of gene expression using microarrays can be used to provide insights into the effects of environmental changes. Such technical developments can be expected to have major impacts on our basic understanding of the degra- dation of aromatic substances in the absence of oxygen, as well as on the ability to manipulate the outcome of natural processes.

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Mägli, A., Wendt, M., Leisinger, T., 1996. Isolation and characterization of Dehalobacterium formicoaceticum gen. nov. sp. nov., a strictly anaerobic bacterium utilizing dichloro- methane as source of carbon and energy. Arch. Microbiol. 166 (2), 101108. Mallick, S., Chakraborty, J., Dutta, T.K., 2011. Role of oxygenases in guiding diverse metabolic pathways in the bacterial degradation of low-molecular- weight polycyclic aromatic hydrocarbons: a review. Crit. Rev. Microbiol. 37 (1), 6490. Available from: https://doi.org/10.3109/1040841X.2010.512268. Marcos, M.S., Lozada, M., Di Marzio, W.D., Dionisi, H.M., 2012. Abundance, dynamics, and biogeographic distribution of seven polycyclic aromatic hydrocarbon dioxygenase gene variants in coastal sediments of Patagonia. Appl. Environ. Microbiol. 78 (5), 15891592. Meagher, R.B., 2000. Phytoremediation of toxic elemental and organic pollutants. Curr. Opin. Plant Biol. 3 (2), 153162. Means, J.C., Wood, S.G., Hassett, J.J., Banwart, W.L., 1980. Sorption of polynuclear aro- matic hydrocarbons by sediments and soils. Environ. Sci. Technol. 14 (12), 15241528. Megharaj, M., Venkateswarlu, K., Naidu, R., 2014. Bioremediation. In: Wexler, P. (Ed.), Encyclopedia of Toxicology, third ed. Academic Press, Oxford, pp. 485489. Miller, C.D., Hall, K., Liang, Y.N., Nieman, K., Sorensen, D., Issa, B., et al., 2004. Isolation and characterization of polycyclic aromatic hydrocarbondegrading Mycobacterium isolates from soil. Microb. Ecol. 48 (2), 230238. Miller, T.R., Heidler, J., Chillrud, S.N., DeLaquil, A., Ritchie, J.C., Mihalic, J.N., et al., 2008. Fate of triclosan and evidence for reductive dechlorination of triclocarban in estuarine sediments. Environ. Sci. Technol. 42 (12), 45704576. Mohn, W.W., Tiedje, J.M., 1992. Microbial reductive dehalogenation. Microbiol. Mol. Biol. Rev 56 (3), 482507. Morris, P.J., Mohn, W.W., Quensen, J.D., Tiedje, J.M., Boyd, S.A., 1992. Establishment of polychlorinated biphenyl-degrading enrichment culture with predominantly meta dechlorination. Appl. Environ. Microbiol 58 (9), 30883094. Nies, L., Vogel, T.M., 1990. Effects of organic substrates on dechlorination of Aroclor 1242 in anaerobic sediments. Appl. Environ. Microbiol 56 (9), 26122617. Payne, R.B., Fagervold, S.K., May, H.D., Sowers, K.R., 2013. Remediation of polychlori- nated biphenyl impacted sediment by concurrent bioaugmentation with anaerobic halor- espiring and aerobic degrading bacteria. Environ. Sci. Technol. 47 (8), 38073815. Perelo, L.W., 2010. In situ and bioremediation of organic pollutants in aquatic sediments. J. Hazard. Mater. 177 (13), 8189. Quensen, J.F., Boyd, S.A., Tiedje, J.M., 1990. Dechlorination of four commercial poly- chlorinated biphenyl mixtures (Aroclors) by anaerobic microorganisms from sediments. Appl. Environ. Microbiol 56 (8), 23602369. Readman, J.W., Mantoura, R.F.C., Rhead, M.M., Brown, L., 1982. Aquatic distribution and heterotrophic degradation of polycyclic aromatic hydrocarbons (PAH) in the Tamar Estuary. Estuar. Coast. Shelf Sci. 14 (4), 369389. Readman, J.W., Fillmann, G., Tolosa, I., Bartocci, J., Villeneuve, J.P., Catinni, C., et al., 2002. Petroleum and PAH contamination of the Black Sea. Mar. Pollut. Bull. 44 (1), 4862. Schink, B., 1992. Syntrophism among prokaryotes. In: Balows, A., Truper, H.G., Dworkin, M., Harder, W., Schleifer, K.H. (Eds.), The Prokaryotes. Springer Verlag, New York, pp. 276299. Scragg, A., 2005. Environmental Biotechnology, second ed. Oxford University Press, New York. Segstro, M.D., Muir, D.C.G., Servo, M.R., Webster, G.R.B., 1995. Long-term fate and bioavailability of sediment associated polychlorinated debenzo-p-dioxins in aquatic mesocoms. Environ. Toxicol. Chem. 14, 17991807. 166 Microbial Communities in Coastal Sediments

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6.1 Introduction Microbial communities comprise vast majority of the earth’s biodiversity and perform significant roles in ecosystem functioning such as organic matter decomposition, biogeochemical cycling of nutrients, and xenobi- otic degradation. This is particularly true for coastal sediment microbial communities as they perform all the above regulatory roles in the land sea interface. Coastal microbial communities encompass a large taxonomic and metabolic diversity that reflects their long history of evolutionary diversification (Logares et al., 2014) and the diversity is pronounced in sediments than water due to large surface area for microorganisms to attach. Hence, the identification and description of microbial biodiversity patterns are important for understanding the biological underpinnings of coastal ecosystem functions. Coastal microbial communities are complex and varying over spatial and temporal scales. Moreover, microbial diversity is derived as the complexity and variability among microorganisms at dif- ferent levels of biological organization such as genetic diversity, species diversity, and , and also the evolutionary and functional processes that link them. Other major factors to be considered are species richness, abundance as well as evenness of distribution that are defined within species diversity. Especially the microbes in the coastal zones are exposed to extreme environmental conditions such as temperature, pH, salinity, and are subjected to various interactions, leading to production of novel metabolites. Much of these remain unexplored due to technological constraints. Hence, unraveling the black box of microbial diversity and characterization of hitherto uncultured microorganisms in coastal sedi- ments so as to obtain a knowledge of their composition and ecological adaptations is important (Köpke et al., 2005). Use of traditional methods depending on the cultivation of microor- ganisms to understand microbial diversity based on physiological and

Microbial Communities in Coastal Sediments © 2021 Elsevier Inc. DOI: https://doi.org/10.1016/B978-0-12-815165-5.00006-6 All rights reserved. 167 168 Microbial Communities in Coastal Sediments biochemical methods is a challenge. This is because of the reason that a vast majority of microbial diversity is unculturable and only 1% of the nat- urally occurring bacteria can be isolated and characterized (Muyzer, 1999a,b). The advent of molecular tools has revolutionized the field of microbial ecology and has led to a wealth of new information on micro- bial diversity and functions in the coastal ecosystems during the past few decades. Understanding the metabolic processes and interactions of coastal sediment microorganisms that shape the microbial community dynamics using advanced high-throughput techniques such as metagenomics has been proven to be powerful in linking the microbial community structure to biogeochemical transformations in the ecosystems (Grossart et al., 2020). Nevertheless, this has led to a surge in microbial community ecol- ogy research leading to disclosure of spatial and temporal patterns of coastal microbial diversity and the dynamics between microorganisms and their surrounding environment. This chapter identifies the weakness of traditional culture-dependent methods and explores the application of novel next-generation technologies that involve culture-independent molecular tools in providing detailed information on coastal microbial diversity and functions.

6.2 Culture-dependent methods: the “great plate count anomaly” Culture-dependent methods involve the viable plate count or most- probable-number techniques that have been/still been used as an inexpen- sive method for quantification of active cells of heterotrophic microbial population in the environment. However, these methods are a challenge to assess the microbial diversity of uncultivable and fastidious species present in low abundance (Bing-Ru et al., 2006). The viable cell counts reported by such culture-dependent methods are several orders of magnitude less than the direct microscopic counts reported in sediments. For example, Torsvik et al. (1998) reported that only 0.1%1.0% of the total bacterial population can be cultured by using the cultivation-dependent procedures and they attributed this to limitations in growth conditions. Most of the cells that appear to be viable cells when visualized microscopically do not form visible colonies on plates. Such difference in cell counts is attributed to factors such as (1) the presence of bacteria in aggregates, (2) selective nature of culture media, (3) the presence of inactive cells, or (4) a considerable portion of the bacteria present in a resting stage (Jannasch and Jones, 1959). To describe Assessment of microbial structure and functions in coastal sediments 169 this phenomenon, Stanley and Konopka (1985) coined the phrase “great plate count anomaly” that represents a vast majority of microorganisms that are active in the natural environment and not cultivable. Moreover, it has also been recognized that the “unculturable” bacterial diversity presents a vast gene pool with ecological significance mediating ecosystem processes (Torsvik et al., 2002). Hence, development of an alternative approach to identify specific microbial populations in their natural environment without cultivation would be considered a revolution in microbial ecology studies (Amann et al., 1995).

6.3 Molecular tools used to examine microbial diversity of coastal sediments Over the last few decades, the understanding of bacterial diversity and dynamics in coastal sediments has significantly increased due to rapid development of culture-independent molecular methods that has provided more detailed information on the phylogeny and distribution of nonculti- vable microorganisms (Amann et al., 1990). The advent of molecular tools revealed an extraordinary diversity of microorganisms in various environ- ments. The genomic age began with the work of Frederick Sanger (the father of genomics) in 1977 for developing techniques to read the nucleo- tides in a strand of DNA and was awarded the Nobel Prize in 1980 (Walker, 2014). The invention of polymerase chain reaction (PCR) by Kary Mullis in early 1980s further revolutionized the approaches of researchers in modern molecular biology and Mullis won the Nobel Prize for this invention in 1993. PCR is a technique that is used to amplify pieces of DNA by several orders of magnitude, up to a million copies of DNA from a single strand. PCR-based molecular methods consist of nucleic acid extraction, amplification of ribosomal DNA (rDNA), and analysis of PCR products by fingerprinting techniques such as denaturing gradient gel electrophore- sis (DGGE) and temperature gradient gel electrophoresis (TGGE), which provide information about the microbial community structure in terms of species richness, evenness, and composition. Another culture-independent DNA fingerprinting method is terminal restriction fragment length poly- morphism (T-RFLP), which has evolved as a more accurate, reliable, speed, and cheaper method of investigating microbial diversity in coastal sediment samples (Thiyagarajan et al., 2010) and enabled scientists to obtain more realistic information about microbes in the environment. 170 Microbial Communities in Coastal Sediments

6.3.1 Gene amplification and sequencing of 16S rDNA The key concept was that organisms can be identified without cultivation and microbial ecology can be studied based on the molecular phylogeny of ribosomal RNA (rRNA) or small subunit rRNA, in particular (Rappé and Giovannoni, 2003). Sequencing of 16S rRNA has represented a fundamen- tal step for identification and classification of bacteria since its advent in 1970s. The first report of phylogenetic composition of bacterial community in coastal marine sediments using molecular tools was by Gray and Herwig (1996). They standardized the protocols for efficient lysis and recovery of DNA from marine sediments. Further microbial DNA was extracted from sediment samples and the 16S rDNA extracted from the samples was ampli- fied using suitable primers using PCR. Gray and Herwig were able to amplify and clone the 16S rDNA from phylogenetically diverse members of bacteria from varied marine sediments. These protocols have enabled microbial ecologists to understand the structure and dynamics of microbial community within the coastal and marine sediments and also understand the fact that vast majority of microbial species in the coastal sediments have not been cultivated yet. However, the cultivation-independent approach of 16S rDNA cloning has aided to elucidate common features within the microbial communities of specific habitats such as coastal benthic environ- ment. The term “species” is substituted with operational taxonomic units (OTUs), which describes those organisms with higher than 97% similarity of 16S rRNA sequence to the same species. This may not always be true. The presence of physiological, genotypic, and phenotypic differences in three species of the genus Bacillus that share .99% similarity in 16S rRNA sequence is a classic example. Phylogenetic analysis of microbial communities in coastal sediments of Puget Sound by amplification of 16S rDNA extracted from sediment sam- ples was done using suitable primers (Gray and Herwig, 1996). The study revealed the presence of diverse population of organisms from the domain bacteria with members of six major lineages represented by alpha, beta, gamma-proteobacteria; clostridia and related organisms; the gram-positive high G 1 C content subdivision; and planctomycetes and related organisms. The microbial community composition in a heavy metalcontaminated marine sediment in Norway investigated using 16S rRNA sequences revealed that the majority of them belonged to the gamma- and delta- proteobacteria and Cytophaga-Flexibacter-Bacteroides bacteria (Gillan et al., 2005). Phylogenetic analysis of full-length 16S rDNA revealed 19 bacterial Assessment of microbial structure and functions in coastal sediments 171 phyla in coastal sediments of Hong Kong, and gamma- and delta- proteobacteria were recorded in all the libraries (Zhang et al., 2007). 16S rRNA gene clone libraries revealed the dominance of gamma proteobacter- ial sequences in varied intertidal sediments (Alonso et al., 2010; Feng et al., 2009). Analysis of microbial community structure by 16S rDNA pyrose- quencing in intertidal sediments exposed to pollution from oil refinery showed a distinct OTUs belonging to Thauera sp., which is known by its ability to metabolize aromatic compounds (Piccini and Garcia-Alonso, 2015). Microbial communities’ studies using 16S rRNA gene amplicon sequencing along the coastline of Puerto Nuevo, Baja California, Mexico revealed that the sediment hosted 500-fold more OTUs than seawater and the phyla found in the sediment were distinct (Ul-Hasan et al., 2019). Although microbial diversity can be studied by PCR-based 16S rDNA cloning methods, the community structure cannot be deduced due to biases during DNA retrieval and amplification (Suzuki and Giovannoni, 1996). Nevertheless, such methods have provided additional DNA sequence information for further design and evaluation of nucleic acid probes so that distinct microbial populations are identified and quantified using quantitative methods such as fluorescence in situ hybridization (FISH) and rRNA slot blot hybridization. Moreover, for studying the diversity and composition of complex microbial communities, compara- tive 16S rRNA sequence analysis and FISH are suggested as key methods (Amann et al., 1995).

6.3.2 Fluorescence in situ hybridization FISH is a qualitative method of nucleic acid hybridization performed on the DNA or RNA extracted from microorganisms and allows examina- tion of specific types or groups of microorganisms in complex environ- mental samples with minimum disruption of the natural microbial community. Using in situ hybridization with rRNA-targeted fluorescent oligonucleotide probes, FISH permits the identification and quantification of individual cells, which has yielded insights into bacterial diversity and community composition in several environments. Since its introduction in the late 1980s, FISH has wide applications in environmental microbiology and considered as a powerful tool for microbial phylogenetic and ecologi- cal studies. The rRNA-targeted approach permits designing of oligonucle- otide probes with specificities that range from the lower species level to higher levels of phyla and domains (Amann and Fuchs, 2008). 172 Microbial Communities in Coastal Sediments

Probe

Sample

Fixation Target (ribosomal RNA)

+ – + + Epifluorescence – + Fixed cells are microscopy – permeabilized + Ribosome – +

Fluorescently labeled Flow cytometry oligonucleotides (probes) Hybridization Quantification

Washing

Hybridized cells Figure 6.1 Steps of FISH using rRNA-targeted nucleotide (Amann and Fuchs, 2008).

Technique: The steps of FISH using rRNA-targeted nucleotide are illustrated in Fig. 6.1. The microbial cells are fixed and incubated with probe during which the labeled oligonucleotide diffuses and forms specific hybrids. Fixing the cells stabilizes the cell morphology and permeabilizes cell membrane for hybridization. After washing of excess probe, the sam- ple is ready for single-cell identification and quantification by epifluores- cence microscopy. Flow cytometry, which is fluorescence-activated cell sorting (FACS), is also used instead of epifluorescence microscopy (Sekar et al., 2004). The first application of FISH to analyze microbial communities in coastal sediments was attempted by Llobet-Brossa et al. (1998) in Wadden Sea sediments of the German North Sea coast by in situ hybridization with group-specific fluorescently labeled, rRNA-targeted oligonucleo- tides. They reported that a large fraction of up to 73% of the DAPI (40,60- diamidino-2-phenylindole)-stained cells were hybridized with the probes and ever since, this technique has been quite frequently used to directly quantify specific communities in sediment. Ravenschlag et al. (2001) used FISH and rRNA slot blot hybridization with 16S rRNA-targeted nucleo- tide probes (bacterial probe EUB 338 and Archaeal probe ARCH915) to investigate the composition of microbial communities in a marine Arctic sediment and found that members of the phylum gamma-proteobacteria Assessment of microbial structure and functions in coastal sediments 173 constituted a significant fraction in the sediment. FISH has been used to study the depth profile of sulfate-reducing bacteria (SRB) in relation to potential electron donors such as acetate and lactate (Llobet-Brossa et al., 2002). Environmentally important anammox bacteria designated as key- stone taxa were detected using FISH in Black Sea sediments and the cell counts of these anammox bacteria were directly correlated to the turnover of ammonium and nitrite (Kuypers et al., 2003). Bacteria of seven phylo- genetic groups were detected in heavy metalcontaminated marine sedi- ments in Norway using FISH and the Desulfosarcina-Desulfococcus represented 6%14% of the DAPI counts (Gillan et al., 2005). The limi- tation of this technique is that FISH does not detect cells with low ribo- some content, which is correlated with microbes of low physiological activity. Hence, the disadvantage of FISH is that it does not detect slow- growing and starving cells (Amann et al., 1995). In addition, FISH detec- tion of microbial communities decreases sharply with increase in sediment depth (Llobet-Brossa et al., 1998)

6.3.3 Terminal restriction fragment length polymorphism T-RFLP is a molecular tool that can be used to estimate the relative abun- dance of dominant bacterial species in coastal sediments and also understand the shift in the pattern of bacterial community structures (Thiyagarajan et al., 2010). For T-RFLP, the community DNA is isolated from coastal sediments and the DNA coding for 16S rRNA is specifically amplified by PCR using primer pairs designed from the conserved region of the gene. After digestion of the PCR product by a restriction enzyme, the length profile of terminal restriction fragment is labeled by a fluorescent dye (Fig. 6.2). The species composition of the bacterial communities is esti- mated by measuring the intensity of fluorescence emission and the ratio of each PCR amplicon indicates the relative abundance of bacterial species. T-RFLP method was used to analyze changes in bacterial community compositions in the coastal surface sediments along a pollution gradient in Hong Kong by Thiyagarajan et al. (2010). They observed that T-RFLP technique is a sensitive and effective method for detecting changes in bac- terial community composition in response to different kinds of environ- mental and anthropogenic disturbances. However, the limitations of this method are that it is a semiquantitative method and does not allow the identification of bacteria in the sediment samples. Hence, for identification of bacteria in the sediments, alternate methods such as DGGE (Muyzer 174 Microbial Communities in Coastal Sediments

Figure 6.2 Workflow of T-RFLP (Thiyagarajan et al., 2010). et al., 1993) need to be done, which allow the extraction and sequence of specific bands resolved on the gel.

6.3.4 Denaturing gradient gel electrophoresis/temperature gradient gel electrophoresis DGGE is a DNA fingerprinting type thought to sample abundant phylo- types ( . 1% of the total) (Pedrós-Alió, 2006). DGGE and TGGE are similar methods for diversity analysis based on separation of PCR- amplified 16S rDNA in a gradient of denaturing agent for DGGE and temperature for TGGE (Muyzer, 1999a,b). The principle of these meth- ods is that upon denaturation, DNA melts and differentially migrates through the polyacrylamide gel based on their melting behavior. This technique is used in environmental microbiology to study microbial com- munity ecology and monitor change in population shifts. As this approach does not provide the phylogenetic information, it can be obtained by excising bands and sequencing the DNA. DGGE can be combined with hybridization using phylogenetic probes or with sequencing to assess the numerically dominating microbes in a community. DGGE was used to detect copper-sensitive bacterial populations in North Sea sediments that were subjected to acute copper exposure, which revealed that DGGE is a valuable tool to assess the effect of pollutants on Assessment of microbial structure and functions in coastal sediments 175 sediment microbial communities (Gillan, 2004). Investigations relating to specific bacterial communities in two tidal flat sediments of German Wadden Sea using rRNA gene PCRDGGE revealed the predominance of Gammaproteobacteria in the upper sediment layers and Chloroflexi in the deeper layers (Webster et al., 2007). DGGE was used to assess the depthwise changes in diversity of hydrocarbon-degrading bacteria in a mangrove sediment using the genes encoding 16S rRNA, BamA, and dsrAB as targets (Andrade et al., 2012). BamA-encoding microorganisms are involved in the degradation of aromatic hydrocarbons and dsrAB is related to SRB (dissimilatory sulfate reduction) and the results revealed a decrease in bacterial abundance and diversity with depth, which might be the reason for persistence of hydrocarbons in deep anoxic layers. DGGE fingerprinting data has been proved to be used as a proxy to confirm the variations in microbial community composition and also has the advantage as a quick and inexpensive method (Cleary et al., 2012). Yao et al. (2017) used PCRDGGE fingerprinting method for examining the effect of heavy metals on microbial community structure in the coastal sediments of Jiaozhou Bay, China and suggested the combination of PCRDGGE with multivariate analysis as an efficient method.

6.4 High-throughput sequencing technologies 6.4.1 Metagenomics: an approach based on small subunit ribosomal RNA The term “microbiome” refers to the complete assemblage of microbes in a discrete habitat. Metagenomic approach is defined as the production and analyses of shotgun genomic data from microbial assemblages, which is now used to analyze microbial community structure in diverse environ- ments like coastal sediments. The term “Metagenomics” coined by Handelsman et al. (1998) is now considered as an advanced technology to analyze extensive data with a wide variety of analysis tools accessible to researchers (Mitchell et al., 2017). Pace et al. (1986) revealed the tech- nique of directly cloning and analyzing DNA from the environment and since then, this approach has been used to unravel the vast “uncultured majority” (Rappé and Giovannoni, 2003) of the microbial communities and establish their roles in biogeochemical cycles. The two major fields of metagenomics are amplicon metagenomics, which is sequencing of librar- ies of a PCR-amplified gene of interest and shotgun metagenomics, which includes screening or sequencing of libraries of randomly isolated 176 Microbial Communities in Coastal Sediments

DNA fragments. The commonly applied sequencing platforms to study microbial diversity in coastal sediments are 454 pyrosequencing and Illumina sequencing systems also referred as next-generation sequencing platforms, which reveal an extensive microbial community composition. The field of environmental “-omics” consisting of metagenomics, metatranscriptomics (sequencing of the expressed genetic material), and metaproteomics (exploration of proteins expressed by microbial commu- nity), which employs molecular biological tools to determine the diversity and function of microbial communities has a key role in establishing the role of microbes in mediating important biogeochemical pathways like elucidation of species involved in phosphorus, sulfur, and nitrogen cycling and also to discover new genes and enzymes of industrial and medical interest (Zarraonaindia et al., 2013). As the microbial community structure varies over spatial and temporal scales, any single technique may not be able to elucidate the complex structure and interaction of microbial com- munities. There is an overarching need to incorporate microbes into bio- geochemical models and adopting an integrated modeling approach to biogeochemistry and environmental genomics data will be a powerful tool to explore the nexus between microbial ecology and geochemistry (Reed et al., 2014). Over the last decades, conventional sequencing of 16S rDNA has revealed a lot of information on the diversity of coastal microbial commu- nities using fingerprinting techniques in combination with clone library construction and Sanger sequencing. However, this often provided insuf- ficient coverage to describe and compare diversity of microbial communi- ties of large sample numbers over varied spatial and temporal scales. Recently, high-throughput sequencing (HTS) technologies and the appli- cation of barcode indexing allow the collection of thousands of sequences from a large number of samples simultaneously. Using HTS technologies, insights into microbial diversity have been advanced. More recently, 454 pyrosequencing and Illumina sequencing techniques have yielded deep insights into the microbial community structure, thus enabling the explo- ration of microbial diversity at an unprecedented scale (Logares et al., 2012). Hence, this approach has been applied recently in a variety of environments for analyzing bacterial and archaeal communities. Generally bacterial communities are dominated by few abundant taxa and many rare taxa and the latter are not sampled adequately by the traditional culture- independent methods. However, the advanced sequencing methods can be powerful tools to address questions about both structure and function Assessment of microbial structure and functions in coastal sediments 177 of rare phylotypes. While the traditional culture-independent methods focused on abundant groups of bacteria, the pyrosequencing data will be used to find the relationships between diversity and abundance of bacterial groups in coastal sediments (Huber et al., 2007). The first study on marine microbial ecology by 454 pyrosequencing of rDNA gene (rDNA ampli- cons) was by Sogin et al. (2006). Most often, metagenomics technique alone fails to provide a highly resolved view of the community structure due to its bias toward sequences of the most abundant taxa. Hence, to have a better understanding of the community structure and functions, metagenomics may be combined with FISH, metatranscriptomics, and metaproteomics (Zarraonaindia et al., 2013). Pyrosequencing-based approach to analyze the 16S rRNA gene of bac- teria in two intertidal sediments of Bohai Bay in China revealed a significant difference in the composition and distribution at phylum, class, and genus levels (Wang et al., 2013). In South China Sea sediments, pyrosequencing of 16S rRNA gene analysis revealed the presence of 9726 OTUs, which include 40 bacterial phyla and among these, 12 phyla were found for the first time by Zhu et al. (2013). Reyes et al. (2017) confirmed the role of SRB belonging to Desulfobacteraceae, Desulfuromonadaceae, and Desulfobulbaceae using real-time quantitative PCR (qPCR) and 16s rRNA pyrosequencing. qPCR and Illumina sequencing of SRB revealed the pre- dominance of Proteobacteria as well as the frequent occurrence of sulfate reduction process in the intertidal sediments of Yangtze Estuary, China (Guo et al., 2018).

6.5 Functional diversity of coastal sediment microbes To identify the microorganisms in coastal sediments that carry out specific microbial processes, or “who is doing what” is a big challenge. One way to address this challenge is to cultivate the isolated microbial strains in the laboratory in a specific growth substrate and assess their physiological and biochemical characteristics, which is time-consuming and laborious. A major limitation in this process is that all the functional microorganisms in the environment cannot be isolated and cultivated in the laboratory. To understand the functional role of sediment archaeal communities in degrading detrital proteins in the sediment, density-gradient centrifugation was used to extract the intact cells from sediment. The genomic DNA of the individual cells sorted with FACS was amplified and identified. The study revealed the presence of two types of peptidases (protein-degrading 178 Microbial Communities in Coastal Sediments enzymes): gingipain and clostripain, which is a novel discovery regarding the role of archaea in protein remineralization in anoxic marine sediments. This finding changed the assumption that bacteria drive the primary remi- neralization of organic matter in the marine sedimentary carbon cycle (Lloyd et al., 2013).

6.5.1 Stable isotope probing Among the limited advanced methods used to identify microbes responsi- ble for particular biogeochemical process, stable isotope probing (SIP) holds considerable promise. Stable isotopes such as carbon and nitrogen have been used extensively to study food-web functioning as well as the flow of energy and matter among organisms (Middelburg, 2014) and also have emerged as a versatile tool to study biogeochemical processes. DNA- based SIP is a method that is used to detect active microbes in the envi- ronment and also link taxonomical identity of microorganisms to specific biogeochemical processes (Radajewski et al., 2000). This technique is used to identify microbes that use a specific growth substrate in environ- ment (Dumont and Murrell, 2005) based on the principle “you are what you eat.” SIP involves exposing an environmental sample to stable isotope like 13C-enriched substrates and identify the microorganisms that have incorporated the 13C substrate. For example, SIP can be used to link active populations of sulfate reducers and methanotrophs to specific bio- geochemical processes by use of substrates 13C-acetate and 13C-methane for the respective group of microorganisms. DNA is isolated from the microbes and subjected to cesium chloride (CsCl) density-gradient centri- fugation with ethidium bromide. The heavy DNA which is 13C-DNA, separates from the light DNA which is 12C-DNA and is visible under ultraviolet illumination. The fraction of microorganisms that contain the 13C-DNA harbors the microorganisms that are active and have incorpo- rated the isotope-labeled substrate into their nucleic acids (Fig. 6.3). A study involving SIP was used to assess the anaerobic carbon mineral- ization pattern in an intertidal estuary (Tamar Estuary, United Kingdom) which revealed that acetate added in a 13C-labeled form was predomi- nantly consumed by SRB and sulfate reduction accounted for 10%60% of carbon mineralization (Boschker et al., 1998). A novel application of DNA-SIP in coastal sediments is to understand the syntrophic associations in anoxic environments (Dumont and Murrell, 2005). Difference in organic substrate utilization pattern by major microbial groups in a vertical Assessment of microbial structure and functions in coastal sediments 179

Figure 6.3 DNA-based stable isotope probing (Dumont and Murrell, 2005). (A) Samples incubated in 13C-labeled substrate, so that the labeled carbon from the sub- strate is incorporated into the biomass of the active microorganisms in the sample. (B) DNA separation by CsCl gradient centrifugation; phylogenetic analysis of sequence data produced by PCR amplification of isolated DNA using specific primers. (c) PLFA (phospho lipid fatty acid) profiles can also reveal the diversity of microbes that incorporated the 13C labeled substrate. profile was detected using SIP of magnetic-bead-captured rRNA in inter- tidal marine sediments of the Oosterschelde Bay in the Netherlands (Miyatake et al., 2014). Investigations on the terminal electron-accepting pathways of acetate-oxidizing bacteria observed that members of the Oceanospirillaceae were identified as the 13C-acetate oxidizers in the sur- face sediment of Aarhus Bay, Denmark. Members of Desulfuromonadales oxidized acetate under iron- and manganese-reducing conditions, whereas members of Gammaproteobacteria oxidized acetate under aerobic and nitrate-reducing conditions (Vandieken and Thamdrup, 2012). In coastal sediments, knowledge on the functional diversity of specific groups of decomposers and mechanisms of organic matter decomposition during carbon cycling is limited. The microbial fate of macrophyte- derived carbon in benthic salt marsh of Southern California was studied by SIP utilizing 13C-labeled lignocellulosic substrate. The study detected 180 Microbial Communities in Coastal Sediments

146 bacterial species and revealed that the predominant groups responsible for utilization of heavy-labeled lignocellulosic substrate in the salt marsh sediments were Desulfosarcina, Spriochaeta, and Kangiella (Darjany et al., 2014). Biogeochemical cycles in the coastal benthic environment are often regulated by hypoxic and anoxic conditions that change periodically in the sediment. The effect of oxygen depletion in the sediments on micro- bial communities and their functions were studied by Coskun et al. 18 18 (2019) using H2 O quantitative SIP, where H2 O is used as a passive 18 tracer. During DNA replication, the oxygen atoms from H2 O are incor- porated into DNA. Further, qPCR combined with HTS of 16S rRNA genes was used to quantify the activity of growing populations among the total microbial communities in the sediment. The findings revealed the predominance of Acidobacteria in the SRB group, thus highlighting the importance of sulfur cycling in the sediments. Glycine betaine (GBT) is the precursor for methane production in coastal marine sediment and 90% of methane production in the coastal eco- systems originates from GBT as well as trimethylamine, and intermediate of GBT through microbial metabolism. To reveal the microorganisms respon- sible for methanogenesis from GBT, a method combining DNA-SIP with metagenomics sequencing and assembly was used to retrieve the complete genome of a novel clostridial bacteria involved in GBT degradation (Jones et al., 2019). In coastal sediments dark carbon fixation (DCF) is a process used by chemoautotrophic microorganisms to synthesize organic molecules from dissolved inorganic carbon. The energy for this process is obtained by the oxidation of various reduced inorganic substrates such as ammonium nitrite, ferrous iron, and sulfide. Global DCF estimated by combining SIP (13C-DIC) with bacterial markers such as PLFA (phospholipid fatty acids) revealed that it was higher in the intertidal sediments than continental shelf- sediments (Vasquez-Cardenas et al., 2020).

6.6 Microbial activity in coastal sediment: study of biogeochemical reaction rates in laboratory microcosms Biogeochemical reactions such as nitrate reduction rate, iron reduction rate, methane production rate, and sulfate reduction rate can be analyzed in coastal sediments as previously explained by Vincent et al. (2017). Sediment samples collected using a core sampler were mixed with an equal volume of sulfate-depleted artificial seawater that has been bubbled with ultrahigh Assessment of microbial structure and functions in coastal sediments 181

Figure 6.4 Analysis of biogeochemical processes in sediments. Photo courtesy: Salom Gnana Thanga Vincent. pure nitrogen gas. A part of the sediment slurry was dispensed into serum vials and sealed with butyl-rubber septa and screw caps and incubated. The biogeochemical rates such as methanogenesis, sulfate reduction, denitrifica- tion, and iron reduction were calculated based on the difference in the con- centration of headspace methane, pore water sulfate, nitrogen, and iron, respectively, before and after incubations (Fig. 6.4).

6.7 Conclusion and future prospects Although microbial communities in the coastal sediments are a dominant driver of various biogeochemical processes, exploration into their struc- tural and functional diversity as well as their ecological determinants has always been a challenge to environmental microbiologists. Advent of novel molecular tools has resulted in surge in microbial community ecol- ogy research in coastal sediments over the past few decades. The “big data” of microbial ecology obtained using advanced molecular methods disclosed the functional traits of microbial communities and elevated our understanding on how microbial communities shape the biogeochemical 182 Microbial Communities in Coastal Sediments cycling patterns in coastal sediments. Also, knowledge on spatial and tem- poral patterns of microbial diversity and their relation to various ecological processes has vastly expanded. Although these innovative discoveries have provided new insights into coastal microbial diversity and functions, much more still remain unexplored. For example, most of the studies pertaining to coastal microbial diversity were reported in “exotic locales” or western world countries. This reveals the presence of large proportions of either unsampled or undersampled coastal sediments around the world where the microbial diversity as well as its associated biogeochemical cycling pat- terns are poorly characterized. Recently marine microorganisms consid- ered to live in “extreme habitats” have been focused for bioprospecting of novel biomolecules with biotechnological applications. Several microbes isolated from coastal environments have proved as important source of biological products and activities with industrial, agricultural, environmen- tal, and pharmaceutical applications. Nevertheless, understanding and exploring the unknown coastal microbial diversity will certainly go a long way to meet the challenges of humanity such as sustainable supply of food and energy, human health, climate change effects, and environmen- tal degradation.

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During the course of a long evolutionary history in a constantly chang- ing environment, microbes in coastal sediments have developed tre- mendous functional resilience. Generally, microorganisms in coastal sediments are characterized by higher genetic diversity and increased than other ecosystems. The structure and functions of microbial communities in coastal sediments are shaped by various factors. The predominant factors are the quality and quantity of organic matter, availability of electron acceptors and donors, the presence of toxic pollutants, and the interactions of microorganisms with other microorganisms as well as with plants and animals. Several studies using model simulations predict that the microbial com- munity structure in coastal sediments will change with global warming and changing nutrient availability. An increase in water temperature due to global warming elevates the biological and chemical reaction rates, for example, enzyme denaturation. In addition, an intensified thermal stratifica- tion of the water column may reduce upwelling, reducing the nutrient supply to the phytoplanktons in the water column. A shift from large- celled phytoplankton such as diatoms to smaller sized picocyanobacteria will alter the quantity and quality of sinking organic matter in the sediments. However, human-induced changes in nutrient cycling are not uni- form around the world due to differences in economic development and governance. Intense nutrient input and a worldwide expansion of hypoxic conditions in the coastal environment of countries with high population density are also expected to have large ramifications for the structure and functions of microorganisms. In places of increased nutri- ent input into coastal ecosystems, eutrophication-induced hypoxia and anoxia stimulate facultative anaerobic and obligate anaerobic microbial activities in the sediments. These processes, on the whole, may lead to increased emission of potent greenhouse gases like nitrous oxide and methane, thus converting coastal ecosystems to net sources of greenhouse gases. However, uncertainty exists on how the anthropogenic changes will reshape microbial communities. It is predicted that by the end of this 1 century, the proton (H ) concentration in the ocean will be twice as

187 188 Appendix 1: Conclusions and future perspectives high as during preindustrial times due to ocean acidification resulting from CO2 absorption. This will result in changes in water and sediment chemistry and biology. In addition, contamination of coastal sediments with persistent organic pollutants is increasing globally. Sediment bacteria capable of degrading these pollutants by utilizing them as nutrient sources are now widely reported. The advent of molecular tools and sophisticated equipment can now aid in exploring the potential of novel microbial communities in coastal sediments and also harnessing their potential for various environmental applications. Index

Note: Page numbers followed by “f” refer to figures.

A B Acetobacterium dehalogenans, 160 Bacterial communities, diversity of, 2030 Acetoclastic methanogenesis, 3233 denitrifying bacteria, 2325 Actinobacteria, 2223 hydrolytic bacteria, 2123 Aerobic respiration, 9091 iron and manganese reducers, 2526 Agricultural and urban runoff, 910 sulfate reducers, 2630 Agriculture, 5254 Bacteroidetes, 1819, 85, 106 Allochthonous eutrophication, 6162 Basal power requirement (BPR), 101 Allochthonous organic matter, 710 Benthic hypoxia, 69 agricultural and urban runoff, 910 Benzoate-degrading bacterium, 152153 transport by rivers, 79 Big data, 181182 Amensalism, 96 Biogeocycling of nutrients, 119133 Ammonium, 104 carbon, 120125 Anaerobic degradation, 149159 greenhouse gas dynamics in coastal 3-chlorobenzoate, 152153, 152f ecosystems, 133139 limitations for, 161 carbon dioxide, 133135 phenols and chlorinated phenols, methane, 135137 151152 nitrous oxide, 138139 polychlorinated biphenyls (PCBs), manganese and iron, 131133 156158 nitrogen, 125129 polychlorinated dibenzo-p-dioxins and sulfur, 129131 dibenzofurans, 158159 Biopolymeric carbon (BPC), 4 polycyclic aromatic hydrocarbons Bioturbation, 9293 (PAHs), 153156, 154f Blue carbon, 123124 Anaerobic microorganisms, 159160 Anammox, 59 C Animal husbandry and marine aquaculture, Carbon cycle, 121122, 122f 5456 Carbon dioxide, 133135 Anoxia in water and sediment, 6871 Carbon disulfide, 130131 Anthropocene, 4748, 5759 Catabolic response profile, 101102 Anthropogenic eutrophication, 67 Chemolithotrophic bacteria, 132 Aquaculture, 5556 Chemolithotrophic denitrifying bacterial Aquatic macrophytes, 136 communities, 127128 Archaeal communities, diversity of, Chlorinated organic compounds, 147 3034 anaerobic transformation of, 151 methanogenic archaea (MA), 3-Chlorobenzoate, 152153, 152f 3234 Clostridia, 27 Atmospheric deposition, 5659 Clostripain, 177178 Atribacteria, 22 Coastal filters, 71 Autochthonous eutrophication, 6162 Coastal microbial communities, 167 Autochthonous organic matter, 57

189 190 Index

Commensalism, 96 Dissolved inorganic carbon (DIC), 34, Competitive substrates, 3233 78, 121 Continental margin sediments, 5 Dissolved inorganic nitrogen (DIN), 59 Copiotrophs, 1819 Dissolved organic carbon (DOC), 3, 1011, Corophium volutator, 103104 8385, 102103, 120121 Crenarchaeota, 27 Dissolved organic matter (DOM), 3, 121 Culture-dependent methods, 168169 Cycloclasticus sp., 155 156 E Cytophaga-Flexibacter-Bacteroides bacteria, Ecological drift, 99 170171 Effluent aura, 6667 Electron acceptors, 161 D changes in availability of, 8894 Dark carbon fixation (DCF), 180 aerobic respiration, 9091 “Dead zone” in the Gulf of Mexico, methanogenesis, 9394 6263, 6869 Mn and Fe reduction, 9293 Degradation Index, 1011 nitrate reduction, 9192 Dehalobacterium formicoaceticum, 160 sulfate reduction, 93 Dehalococcoides,27 vertical organization of, 90f Dehalogenation reactions, 150151 Electron donors, role of, 9495 Dehalorespiration, 159160 Environmental variables and factors Deltaproteobacteria, 2527, 29, 106 regulating microbial structure and Denaturing gradient gel electrophoresis functions, 79 (DGGE), 169, 174175 biological factors, 95105 Denitrification, 23, 59, 72, 92, 104, geological factors, 8285 126127, 138139 sediment depth, 8385 Denitrifiers, 24 sediment granulometry, 8283 Denitrifying bacteria, 2325 hydrological factors, 85 Desulfitobacterium frappieri, 160 natural and anthropogenic disturbances, Desulfobacteraceae, 2627, 177 106 Desulfobacterales, 27 nutritional factors, 105106 Desulfobulbaceae, 2627, 177 physicochemical factors, 8595 Desulfomonile tiedjeii, 152153 electron acceptors, changes in Desulfosporosinus,30 availability of, 8894 Desulfovibrio,30 electron donors, role of, 9495 Desulfuromonadaceae, 177 pH, 86 Dichloro-diphenyl-trichloroethane (DDT), pore water chemistry/presence of 147 nutrients or chemicals, 87 Dimethyl disulfide (DMDS), 130131 redox potential, 8788 Dimethyl sulfide (DMS), 130131 salinity, 8687 Dimethyl sulfoniopropionate (DMSP), presence of contaminants/toxic 130131 substances, 107108 Dimethyl sulfoxide (DMSO), 130131 spatial and temporal heterogeneity, Dioxins, 147 8082 Dissimilatory nitrate reduction, 2324 trophic interactions, 9598 Dissimilatory nitrate reduction to bioturbation and ventilation, ammonium (DNRA), 59, 127128 103104 Dissimilatory sulfate reduction, 93 ecological coherence, 100101 Index 191

evolutionary mechanisms and H diversification, 98100 High-throughput sequencing (HTS) microbial characteristics, 101103 technologies, 175177 plant interactions, 104105 metagenomics, 175177 syntrophy and interspecies hydrogen Horizontal gene transfer (HGT), 99100 transfer, 9698 Hydrocarbon-degrading community, 108 Epsilonproteobacteria, 2729 4-Hydroxybenzoate, 151 Euryarchaeota, 27 Hydrogen sulfide, 130131 Eutrophication, 6567, 72, 119 Hydrolytic bacteria, 2123 and consequences for ecology, 6168 Hypernutrification, effect of, 6173 induced changes in sediment microbial eutrophication and consequences for communities, 7173 ecology, 6168 eutrophication-induced changes in F sediment microbial communities, Fermentation, 1213 71 73 Firmicutes, 27 hypoxia and anoxia in water and Fish farming, 5556 sediment, 68 71 Fluorescence-activated cell sorting (FACS), Hypoxia and anoxia in water and 172, 177178 sediment, 68 71 Fluorescence in situ hybridization (FISH), 171173, 172f I Fossil fuel burning and atmospheric Intermediary ecosystem metabolism, 9798 deposition, 5659 International Geosphere-Biosphere Functional diversity of coastal sediment Programme (IGBP), 5051 microbes, 177180 Interspecies hydrogen transfer, 9698 stable isotope probing, 178180 Iron and manganese reducers, 2526, 9293 G Gammaproteobacteria, 2425, 8587, K 106, 174175 K-strategist life histories, 105 Gemmatimonadetes, 22 Gene amplification and sequencing of 16S rDNA, 170171 L Labile organic matter (LOM), 4, 1011 Geobacteraceae, 25 LOICZ (Land-Ocean Interactions in the Geobacter metallireducens, 132 Gingipain, 177178 Coastal Zone), 50 51 Global carbon cycle, 121122, 122f Glycine betaine (GBT), 180 M Great plate count anomaly, 168169 Magnetospirillum, 152153 Greenhouse gas (GHG) emissions, Manganese, 9293, 131133 119120 Manganese and iron reduction, 2526, Greenhouse gas dynamics in coastal 9293 ecosystems, 133139 Marine aquaculture, 5456 carbon dioxide, 133135 Marine Benthic Group B/Deep Sea methane, 135137 Archaeal Group (MBG-B/DSAG), nitrous oxide, 138139 3132 192 Index

Marine Group I (MG-1), 3031 Microbial nitrate reduction, 24 mcrA gene, 32 Microbiome, 175176 Meat production, 55 Mineralization, 119 Metagenomics, 175177 Miscellaneous Crenarchaeotal group Methane, 135137 (MCG), 3031 Methane thiol (MeSH), 130131 Molecular tools used to examine microbial Methanobacteriales, 32 diversity of coastal sediments, Methanocellales, 32 169175 Methanococcales, 32 denaturing gradient gel electrophoresis Methanococcoides,3334 (DGGE), 174175 Methanogenesis, 32, 9394, 135136 fluorescence in situ hybridization (FISH), Methanogenic archaea (MA), 3234, 171173, 172f 8687 future prospects, 181182 Methanogenic microorganisms, 135136 gene amplification and sequencing of Methanogenium,3233 16S rDNA, 170171 Methanogens, 135136 temperature gradient gel electrophoresis Methanomassisilicoccales, 32 (TGGE), 174175 Methanomicrobiales, 3234 terminal restriction fragment length Methanopyrales, 32 polymorphism (T-RFLP), Methanosaeta,3234 173174, 174f Methanosarcina,3334 Monomeric sugars, 11 Methanosarcinales,3233 Methanoseta concii,32 33 N 4-Methylbenzoate, 152153 Nereis virens, 103104 Methyl-coenzyme M reductase (MCR), 32 Niche overlap hypothesis, 1819, 85 Microbe-mediated carbon cycle, 120125, Nitrate reduction, 9192 124f Nitrification, 104, 138139 Microbe-mediated iron cycle, 131133, Nitrogen-based trace gases, 5657 132f Nitrous oxide, 119120, 138139 Microbe-mediated nitrogen cycle, Noncompetitive substrates, 135136 125129, 126f Nutrients, biogeocycling of, 119133 Microbe-mediated sulfur cycle, 129131, carbon, 120125 131f greenhouse gas dynamics in coastal Microbial activity in coastal sediment, ecosystems, 133139 180181 carbon dioxide, 133135 Microbial biomass carbon (MBC), 4 methane, 135137 Microbial degradation of organic matter, nitrous oxide, 138139 1217 manganese and iron, 131133 Microbial diversity and ecology in coastal nitrogen, 125129 sediments, 1930 sulfur, 129131 bacterial communities, diversity of, Nutrient sources of coastal sediments, 47, 2030 5059 denitrifying bacteria, 2325 agriculture, 5254 hydrolytic bacteria, 2123 animal husbandry and marine iron and manganese reducers, 2526 aquaculture, 5456 sulfate reducers, 2630 fossil fuel burning and atmospheric Microbial mats, 18 deposition, 5659 Index 193

hypernutrification, effect of, 6173 P eutrophication and consequences for Particulate inorganic carbon (PIC), 78 ecology, 6168 Particulate organic carbon (POC), 3, 15 eutrophication-induced changes in Particulate organic matter (POM), 3 sediment microbial communities, 2,3,4,5,6-Pentachlorobiphenyls, 157 7173 Persistent organic pollutants (POPs), hypoxia and anoxia in water and 147149 sediment, 6871 anaerobic degradation, limitations for, nutrient enrichment, 5960 161 anaerobic degradation and pathways, O 149159 Odum’s theory of ecosystem succession, 3-chlorobenzoate, 152 153, 152f 1617 phenols and chlorinated phenols, Oligotrophs, 105 151 152 Operational taxonomic units (OTUs), 170 polychlorinated biphenyls (PCBs), Organic carbon (OC), 13 156 158 Organic matter (OM), 14, 119121 polychlorinated dibenzo-p-dioxins anaerobic degradation, 14f and dibenzofurans, 158 159 autochthonous production, 6f polycyclic aromatic hydrocarbons in designing sediment microbial (PAHs), 153 156, 154f communities, 1719 anaerobic microorganisms, 159 160 microbial degradation of, 1217, 13f future prospects, 161 162 sedimentary organic matter, types of, Phenols and chlorinated phenols, 151 152 24 Phosphorus, 63 biopolymeric carbon (BPC), 4 Polychlorinated aromatic hydrocarbons, dissolved inorganic carbon (DIC), 147 148 34 Polychlorinated biphenyls (PCBs), dissolved organic matter (DOM)/ 147 148, 156 158 dissolved organic carbon (DOC), 3 Polychlorinated dibenzo-p-dioxins and labile organic matter carbon dibenzofurans (PCDD/Fs), (LOM-C), 4 158 159 microbial biomass carbon (MBC), 4 Polycyclic aromatic hydrocarbons (PAHs), particulate organic matter (POM)/ 147 148, 153 156, 154f particulate organic carbon Porphyra seaweeds, 65 66 (POC), 3 Proteobacteria, 24 25, 29 refractory organic matter, 4 Protozoan predation, 98 total organic matter (TOM)/total Pyrosequencing-based approach, 177 organic carbon (TOC), 23 types of, 2f Q Organic matter, source of, 510 Quantitative PCR (qPCR), 177 allochthonous organic matter, 710 agricultural and urban runoff, 910 transport by rivers, 79 R Redox potential, 8788 autochthonous organic matter, 5 7 Organic matter quality indices, 1012 Reduction oxidation (redox) state, 87 88 Reductive dehalogenation, 150151 Oxidationreduction reactions, 15 Refractory organic matter, 4 Oxygen, 120 194 Index

Reoxidation, 130 Sulfatemethane transition (SMT) zone, Rhodopseudomonas, 160 17 Rhodospirillum, 160 Sulfate reducers, 2630 Ribosomal DNA (rDNA), 169 Sulfate-reducing bacteria (SRB), 2627, Ribosomal RNA (rRNA), 170171 30, 85, 93, 130, 172173 Sulfate reduction, 93 S Syntrophism, 97f Syntrophy and interspecies hydrogen Salinity, 8687 transfer, 9698 Seagrasses, 67 Sedimentary organic matter, types of, 24 biopolymeric carbon (BPC), 4 T dissolved inorganic carbon (DIC), 34 Temperature gradient gel electrophoresis dissolved organic matter (DOM)/ (TGGE), 169, 174175 dissolved organic carbon (DOC), 3 Terminal electron acceptors (TEAs), 120, labile organic matter carbon (LOM-C), 131132 4 Terminal restriction fragment length microbial biomass carbon (MBC), 4 polymorphism (T-RFLP), 169, particulate organic matter (POM)/ 173174 particulate organic carbon (POC), 3 workflow of, 174f refractory organic matter, 4 1,2,3,4-Tetrachlorodibenzo(p)dioxin, total organic matter (TOM)/total 156157 organic carbon (TOC), 23 Thauera sp., 170171 Sediment depth, 8385 Total organic carbon (TOC), 23 decrease in lability of organic matter Total organic matter (TOM), 23 with depth, 8485 Transport by rivers, 79 shift in substrate availability with depth, Trichlorosan (TCS), 152 8384 Triclocarban (TCC), 152 Sediment granulometry, 8283 Sediment microbial communities U eutrophication-induced changes in, Ulva prolifera,6566 7173 organic matter in designing, 1719 Sediments, quality of organic matter in, V 1012 Volatile fatty acids (VFAs), 15, 27 organic matter quality indices, 1012 Sheltered coastal systems, 61 62 Z Spatial and temporal heterogeneity, 8082 “Zymogenous”,1819, 105 Stable isotope probing, 178180