Characterization of Organic Matter

In chapter 8, a general overview was provided on the dominant sources of organic matter in estuarine systems. In general, estuarine organic matter is derived from a multitude of natural and anthropogenic allochthonous and autochthonous sources that originate across a freshwater to seawater continuum. Knowledge of sources, reactivity, and fate of organic matter are critical in understanding the role of estuarine and coastal systems in global biogeochemical cycles (Simoneit, 1978; Hedges and Keil, 1995; Bianchi and Canuel, 2001). Due to a wide diversity of organic matter sources and the dynamic mixing that occurs in estuarine systems, it remains a significant challenge in determining the relative importance of these source inputs to biogeochemical cycling in the water column of sediments. Temporal and spatial variability in organic matter inputs adds further to the complexity in understanding these environments. In recent years there have been significant improvements in our ability to distinguish between organic matter sources in estuaries using tools such as elemental, isotopic (bulk and compound/class specific), and chemical biomarker methods. This chapter will provide a general overview of the biochemistry of dominant organic compounds in organic matter and the techniques used to distinguish them in estuarine systems. The abundance and ratios of important elements in biological cycles (e.g., C, H, N, O, S, and P) provide the basic foundation of information on organic matter cycling. For example, concentrations of total organic carbon (TOC) provide the most important indicator of organic matter since approximately 50% of most organic matter consists of C. As discussed in chapter 8, TOC in estuaries is derived from a broad spectrum of sources with very different structural properties and decay rates. Consequently, while TOC provides essential information on spatial and temporal dynamics of organic matter it lacks any specificity to source or age of the material. When bulk C information is combined with additional elemental information, as in the case of the C-to-N ratio, basic source information can be inferred about algal and terrestrial source materials (see review, Meyers, 1997). The broad range of C:N ratios across divergent sources of organic matter in the biosphere demonstrate how such a ratio can provide an initial proxy for determining source information.

Organic Matter Cycling

In this chapter the general processes involved in controlling production and transformation of organic matter will be discussed as well as some of the associated stoichiometric changes of a few key biological elements (e.g., C, N, P, S). Stoichiometry is defined as the mass balance of chemical reactions as they relate to the law of definite proportions and conservation of mass (Sterner and Elser, 2002). For example, if we examine the average atomic ratios of C, N, and P in phytoplankton we see a relatively consistent ratio of 106:16:1 in most marine species. This is perhaps the best example of applied stoichiometric principles in natural ecosystems and is derived from the classic work of Alfred C. Redfield (1890–1983) (Redfield, 1958; Redfield et al., 1963). More specifically, Redfield compared the ratios of C, N, and P of dissolved nutrients in marine waters to that of suspended marine particulate matter (seston) (essentially phytoplankton) and found straight lines with equal slopes (figure 8.1; Redfield et al., 1963). This relationship suggested that marine biota were critical in determining the chemistry of the world ocean, clearly one of the most important historical findings linking chemical and biological oceanography (Falkowski, 2000). Moreover, the Redfield ratio has been further validated with recent data using improved analytical techniques (Karl et al., 1993; Hoppema and Goeyens, 1999). Other work has shown that there are predictable deviations from the Redfield ratio across a freshwater to open ocean marine gradient (figure 8.2; Downing, 1997). For example, N-to-P ratios in estuaries have commonly been shown to be lower and/or higher than the predicted Redfield ratio because of denitrification and anthropogenic nutrient enrichment processes, respectively. Inputs of vascular plant organic matter (e.g., mangroves, salt marshes, seagrasses) to estuarine systems presents another problem in causing deviations of C:N:P from the Redfield ratio. Vascular plants have been shown to deviate from this ratio in part because of relatively high amounts of C and N compared to algae due to a higher abundance of structural support molecules (e.g., cellulose, lignin) and defense antiherbivory (secondary) compounds (e.g., tannins), respectively (Vitousek et al., 1988).

Origin and Geomorphology

Geologically speaking, estuaries are ephemeral features of the coasts. Upon formation, most begin to fill in with sediments and, in the absence of sea level changes, would have life spans of only a few thousand to tens of thousands of years (Emery and Uchupi, 1972; Schubel, 1972; Schubel and Hirschberg, 1978). Estuaries have been part of the geologic record for at least the past 200 million years (My) BP (before present; Williams, 1960; Clauzon, 1973). However, modern estuaries are recent features that only formed over the past 5000 to 6000 years during the stable interglacial period of the middle to late Holocene epoch (0–10,000 y BP), which followed an extensive rise in sea level at the end of the Pleistocene epoch (1.8 My to 10,000 y BP; Nichols and Biggs, 1985). There is general agreement that four major glaciation to interglacial periods occurred during the Pleistocene. It has been suggested that sea level was reduced from a maximum of about 80 m above sea level during the Aftoninan interglacial to 100 m below sea level during the Wisconsin, some 15,000 to 18,000 y BP (figure 2.1; Fairbridge, 1961). This lowest sea level phase is referred to as low stand and is usually determined by uncovering the oldest drowned shorelines along continental margins (Davis, 1985, 1996); conversely, the highest sea level phase is referred to as high stand. It is generally accepted that low-stand depth is between 130 and 150 m below present sea level and that sea level rose at a fairly constant rate until about 6000 to 7000 y BP (Belknap and Kraft, 1977). A sea level rise of approximately 10 mm y−1 during this period resulted in many coastal plains being inundated with water and a displacement of the shoreline. The phenomenon of rising (transgression) and falling (regression) sea level over time is referred to as eustacy (Suess, 1906). When examining a simplified sea level curve, we find that the rate of change during the Holocene is fairly representative of the Gulf of Mexico and much of the U.S. Atlantic coastline (Curray, 1965).

Estuarine–Coastal Interactions

The coastal ocean is a dynamic region where the rivers, estuaries, ocean, land, and the atmosphere interact (Walsh, 1988; Mantoura et al., 1991; Alongi, 1998; Wollast, 1998). Coastlines extend over an estimated 350,000 km worldwide, and the coastal ocean, typically defined as a region that extends from the high water mark to the shelf break (figure 16.1; Alongi, 1998), covers approximately 7% (26 × 106 km2) of the surface global ocean (Gattuso et al., 1998). Although relatively small in area, this highly productive region (30% of the total net oceanic productivity) supports as much as 90% of the global fish catch (Holligan, 1992). In recent years, the coastal ocean has been recognized for its global importance with both national and international programs such as the Land–Ocean Interactions in the Coastal Zone (LOICZ) program, a subprogram of the International Global Change Program (IGBP) started in 1993 (Pernetta and Milliman, 1995), the European Union coastal core project (European Land–Ocean Interaction Studies, ELOISE) (Cadée et al., 1994), and in the U.S. Shelf Edge Exchange Processes Program (SEEP I and SEEP II) (Walsh et al., 1988; Anderson et al., 1994), the Coastal Ocean Processes (CoOP) program, Ocean Margins Program (OMP), and Land–Margin Ecosystem Research (LMER), to name a few. SEEP I and SEEP II were designed to test the Walsh et al. (1985) hypothesis that increased anthropogenic nutrient supply to the coastal ocean would result in enhanced burial of organic matter in continental margins due to higher offshore export of new productivity in the nearshore waters. While the hypothesis of offshore transport and burial was shown to be valid along certain regions of the eastern U.S. coast, other regions showed a more along-shelf transport (Walsh, 1994). More recent work in the OMP, which revisited some of the objectives of SEEP I and SEEP II, found that the accumulation of organic matter in upper slope sediments was only <1% of the total primary production in the entire continental margin of North Carolina (DeMaster et al., 2002). There are many factors that will ultimately determine if and how much nearshore production is exported offshore from the coastal ocean.

Trace Metal Cycling

Like many other elements, natural background levels of trace elements exist in crustal rocks, such as shales, sandstones, and metamorphic and igneous rocks (Benjamin and Honeyman, 2000). In particular, the majority of trace metals are derived from igneous rocks, simply based on the relative fraction of igneous rocks in comparison with sedimentary and metamorphic rocks in the Earth’s crust. The release of trace metals from crustal sources is largely controlled by the natural forces of physical and chemical weathering of rocks, notwithstanding large-scale anthropogenic disturbances such as mining, construction, and coal burning (release of fly ash). As discussed later in the chapter, adjustments can be made for anthropogenic loading to different ecosystems based on an enrichment factor which compares metal concentrations in the ecosphere to average crustal composition. Biological effects of weathering, such as plant root growth and organic acid release associated with respiration also contribute to these weathering processes. As some trace metals are more volatile than others, release due to volcanic activity represents another source of metals with such properties (e.g., Pb, Cd, As, and Hg). Just as Goldschmidt (1954) grouped elements (e.g., siderophiles, chalcophiles, lithophiles, andatomophiles) based on similarities in geochemical properties, trace metals also represent a group of elements with similar chemical properties. One particularly important distinguishing feature of these elements is their ability to bond reversibly to a broad spectrum of compounds (Benjamin and Honeyman, 2000). Thus, the major inputs of trace metals to estuaries are derived from riverine, atmospheric, and anthropogenic sources. Although trace elements typically occur at concentrations of less than 1 ppb (part per billion) (or μg L−1, also reported in molar units), these elements are important in estuaries because of their toxic effects, as well as their importance as micronutrients for many organisms. The fate and transport of trace elements in estuaries are controlled by a variety of factors ranging from redox, ionic strength, abundance of adsorbing surfaces, and pH, just to name a few (Wen et al., 1999).

Carbon Cycle

Carbon is the key element of life on Earth and exists in more than a million compounds (Holmén, 2000; Berner, 2004). The unique covalent long-chained and aromatic carbon compounds form the basis of organic chemistry and the “roadmap” for understanding life from the cellular to the ecosystem level. The oxidation states of C atoms range from +IV to −IV; methane (CH4) is the most reduced form of C (−IV), with CO2 and other carbonate forms existing in the most oxidized state (+IV). The major reservoirs of C are stored in the Earth’s crust, with much of it as inorganic carbonate and the remaining as organic C (e.g., kerogen) (figure 13.1; Sundquist, 1993). The global C cycle can be divided into short- and long-term cycles based on the vast differences in the turnover times of different C pools (Berner, 2004). The carbonate reservoir can be divided into two primary subreservoirs: (1) dissolved inorganic carbon (DIC) in the ocean (H2CO3, HCO3−, and CO32−), and (2) solid carbonate minerals [CaCO3, CaMg(CO3)2, and FeCO3] (Holmén, 2000). While the global C cycle is quite complex, it is perhaps the best understood of all the bioactive element cycles. In fact, there have been numerous review papers on this cycle (e.g., Keeling, 1973; Degens et al., 1984; Siegenthaler and Sarmiento, 1993; Sundquist, 1993; Schimel et al., 1995; Holmén, 2000). Much of the interest in the global C cycle in recent years stems from linkages with environmental issues concerning carbon-based greenhouse gases (e.g., CO2 and CH4) and their role in global climate change (Dickinson and Cicerone, 1986). As described in chapter 8, short-term controls on the C cycle are largely a function of the uptake of inorganic C by autotrophs to fuel fixation in photosynthesis, and the utilization of organic carbon as a food resource by heterotrophs recycling inorganic C back into the system. This short-term cycle, which allows for the transfer of C between the lithosphere, hydrosphere, biosphere, and atmosphere over periods of days to thousands of years, is relatively short in comparison to the more than 4 billion year age of the Earth.

Sources and Distribution of Sediments

The uplift of rocks above sea level on the Earth’s surface over geological time, produces rock material that can be altered into soils and sediments by weathering processes. Over geological time, a fraction of sediments can be sequestered for storage in the ocean basins—with most of it stored in the coastal margin. However, much of this material is modified via processing in large river estuarine systems which can ultimately affect the long-term fate of these terrigenous materials. Sediments produced from weathering of igneous, metamorphic, and sedimentary rocks are principally transported to the oceans through river systems of the world. The major routes of sediment transport from land to the open ocean can simply be illustrated through the following sequence: streams, rivers, estuaries, shallow coastal waters, canyons, and the abyssal ocean. It should be noted that significant and long-term storage occurs in river valleys and floodplains (Meade, 1996). Submarine canyons are also thought to be temporary storage sites for land-derived sediments; however, episodic events such as turbidity currents and mud slides can move these sediments from canyons to the abyssal ocean (more details on coastal margin transport to the deep ocean are provided in chapter 16). The annual sediment flux from rivers to the global ocean is estimated to range from 18 to 24 × 109 metric tonnes (Milliman and Syvitski, 1992). Conversely, estuaries will eventually fill-in with fluvial inputs of sediments over time, and ultimately reach an equilibrium whereby export and import of sediment supply are balanced (Meade, 1969). For example, recent studies have shown that sediment accumulation in the Hudson River estuary, both short (Olsen et al., 1978) and long term (Peteet and Wong, 2000), is in equilibrium with sea level rise. More specifically, it is believed that river flow controls the direction of sediment flux in the Hudson, while variations in spring-neap tidal amplitude control the magnitude (Geyer et al., 2001). Weathering is typically separated into two categories: physical and chemical. Physical weathering involves the fragmentation of parent rock materials and minerals through processes such as freezing, thawing, heating, cooling, and bioturbation (e.g., endolithic algae, fungi, plant roots, and earthworms).

Phosphorus and Silica Cycles

Phosphorus (P) is one of the most well-studied nutrients in aquatic ecosystems because of its role in limiting primary production on ecological and geological timescales (van Capellen and Berner, 1989; Holland, 1994; Tyrell, 1999; van Cappellen and Ingall, 1996). Other key linkages to biological systems include the role of P as an essential constituent of genetic material (RNA and DNA) and cellular membranes (phospholipids), as well as in energy-transforming molecules (e.g., ATP, etc.). Consequently, marine P has received considerable attention in recent decades, with particular emphasis on source and sink terms in budgets (Froelich et al., 1982; Meybeck, 1982; Ruttenberg, 1993; Sutula et al., 2004). Excessive loading of N to estuarine waters can result in P limitation in systems that are generally considered to be N limited. In such cases where primary production is limited by P, N:P ratios are expected to exceed the Redfield value of 16:1 but can be replenished by sediment efflux of P due to redox changes. For example, after the initial N loading of a system there will be an increase in primary production, which can cause the system to become P limited. Then, the phytodetritus from these early stages of N loading can be remineralized in sediments resulting in anoxic conditions in surface sediments, which can then enhance P release from sediments to the overlying waters where primary production is once again enhanced. Evidence for the role of sediment-derived P on primary production in estuaries with high N loading has been shown to occur particularly in shallow water systems (Timmons and Price, 1996; Cerco and Seitzinger, 1997). On the other hand, many coastal areas have also been subjected to high P loading from anthropogenic sources, where in some cases inputs of P are 10 to 100 times greater than in preindustrial times (Caraco et al., 1993). In many cases, P and N loading to estuarine systems will occur simultaneously and decoupling or isolating their individual effects can be difficult (e.g., HELCOM, 2001). The cycling and availability of P in estuaries is largely dependent upon P speciation.

Dissolved Gases in Water

Dissolved gases are critically important in many of the biogeochemical cycles of estuaries and coastal waters. However, only recently have there been large-scale collaborative efforts addressing the importance of coupling between estuaries and the atmosphere. For example, the Biogas Transfer in Estuaries (BIOGEST) project, which began in 1996, was focused on determining the distribution of biogases [CO2, CH4, CO, non-methane hydrocarbons, N2O, dimethyl sulfide (DMS), carbonyl sulfide (COS), volatile halogenated organic compounds, and some biogenic volatile metals] in European estuaries and their impact on global budgets (Frankignoulle and Middelburg, 2002). The role of the estuaries and other coastal ocean environments as global sources and/or sinks of key greenhouse gases, like CO2, have also been a subject of intense interest in recent years (Frankignoulle et al., 1996; Cai and Wang, 1998; Raymond et al., 1997, 2000; Cai, 2003; Wang and Cai, 2004). Similarly, O2 transfer across the air–water interface is critical for the survival of most aquatic organisms. Unfortunately, many estuaries around the world are currently undergoing eutrophication, which commonly results in low O2 concentrations (or hypoxic ≤ 2 mg L−1), due to excessive nutrient loading in these systems (Rabalais and Turner, 2001; Rabalais and Nixon, 2002). To understand how gases are transferred across the air–water boundary we will first examine the dominant atmospheric gases and physical parameters that control their transport and solubility in natural waters. The atmosphere is also composed of aerosols, which are defined as condensed phases of solid or liquid particles, suspended in state, that have stability to gravitational separation over a period of observation (Charlson, 2000). Chemical composition and speciation in atmospheric aerosols is important to understanding their behavior after deposition, and is strongly linked with the dominant sources of aerosols (e.g., windblown dust, seasalt, combustion). The importance of aerosol deposition to estuaries and coastal waters, via precipitation (rain and snow) and/or dry particle deposition, has received considerable attention in recent years. For example, dry and wet deposition of nutrients (Paerl et al., 2002; Pollman et al., 2002) and metal contaminants (Siefert et al., 1998; Guentzel et al., 2001) has proven to be significant in biogeochemical budgets in wetlands and estuaries.

Nitrogen Cycle

Elemental nitrogen (N2) makes up 80% of the atmosphere (by volume) and represents the dominant form of atmospheric nitrogen gas. Despite its high atmospheric abundance, N2 is generally nonreactive, due to strong triple bonding between the N atoms, making much of this N2 pool unavailable to organisms. In fact, only 2% of this N2 pool is believed to be available to organisms at any given time (Galloway, 1998). Consequently, N2 must be “fixed” into ionic forms such as NH4+ before it can be used by plants. Since N is essential for the synthesis of amino acids and proteins and because it is often in low concentrations, N is usually considered to be limiting to organisms in many ecosystems. Nitrogen has five valence electrons and can occur in a broad range of oxidation states that range from +V to -III, with NO3− and NH4+ being the most oxidized and reduced forms, respectively. Some of the most common N compounds that exist in nature, along with their boiling points, ΔH0, and ΔG0, are shown in table 10.1 (Jaffe, 2000); these thermodynamic data can be used to calculate equilibrium concentrations. Fluxes in the global N cycle have been seriously altered by anthropogenic activities (Vitousek et al., 1997; Galloway et al., 2004). For example, fluxes of many nitrogen oxides, which are largely derived from burning fossil fuels, have increased significantly in the atmosphere resulting in photochemical smog and acid precipitation (table 10.2; Jaffe, 2000). Similarly, the advent of artificial N fertilizers (e.g., the Haber–Bosch process, where N2 is fixed to NH3 by industrial processes), which were developed to compensate for the general nonavailability of N2 to most agricultural crops, has resulted in increased N loading from soils and sewage to rivers and estuaries around the world, and considerable eutrophication problems in these aquatic ecosystems. For example, biological N2 fixation accounted for a major fraction of newly fixed N before the 1800s (∼90–130 Tg N y−1) (Galloway et al., 1995).