See discussions, st ats, and author pr ofiles f or this public ation at : https:www .researchgate.ne tpublic ation278651309 [616261]

See discussions, st ats, and author pr ofiles f or this public ation at : https://www .researchgate.ne t/public ation/278651309
Dissolved Organic Matter in Natu ral Waters
Chapt er · Januar y 2013
DOI: 10.1007/978-3-642-32223-5_1
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Khan M. G. Most ofa
Tianjin Univ ersity
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Cong-Qiang Liu
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11 Introduction
Organic matter (OM) in water is composed of two major fractions: dissolved
and non-dissolved, defined on the basis of the isolation technique using filters (0.1–0.7 μm). Dissolved organic matter (DOM) is the fraction of organic sub-
stances that passes the filter, while particulate organic matter (POM) remains on the filter (Danielsson 1982; Kennedy et al. 1974; Liu et al. 2007; Mostofa et al. 2009a). DOM is generally originated from three major sources: (i) allochthonous Dissolved Organic Matter in Natural Waters
K. M. G. Mostofa et al. (eds.), Photobiogeochemistry of Organic Matter,
Environmental Science and Engineering, DOI: 10.1007/978-3-642-32223-5_1, © Springer-Verlag Berlin Heidelberg 2013Khan M. G. Mostofa, Cong-qiang Liu, M. Abdul Mottaleb, Guojiang Wan,
Hiroshi Ogawa, Davide Vione, Takahito Yoshioka and Fengchang Wu
K. M. G. Mostofa (*) · C. Q. Liu · G. Wan
State Key Laboratory of Environmental Geochemistry, Institute of Geochemistry, Chinese Academy of Sciences, Guiyang 550002, Chinae-mail: [anonimizat]
M. A. Mottaleb
Center for Innovation and Entrepreneurship (CIE), Department of Chemistry/Physics, Northwest Missouri State University, 800 University Drive, Maryville, MO 64468, USA
H. Ogawa
Atmospheric and Ocean Research Institute, The University of Tokyo, 1-15-1, Minamidai,
Nakano, 164-8639 Tokyo, Japan
D. Vione
Dipartimento di Chimica Analitica, University of Turin, I-10125 Turin, Italy
Centro Interdipartimentale NatRisk, I-10095 Grugliasco (TO), Italy
T. Yoshioka
Field Science Education and Research Center, Kyoto University, Kitashirakawa Oiwake-cho,
Sakyo-ku, Kyoto 606-8502, Japan
F. C. Wu
State Environmental Protection Key Laboratory of Lake Pollution Control, Chinese Research Academy of Environmental Sciences, Chaoyang 100012, China

2 K. M. G. Mostofa et al.
or terrestrial material from soils, (ii) autochthonous or surface water-derived of
algal or phytoplankton origin, and (iii) syhthetic organic substances of man-made or industrial origin. DOM in natural waters is composed of a heterogeneous mix-ture of organic compounds with molecular weights ranging from less than 100 to over 300,000 Daltons (Hayase and Tsubota 1985; Thurman 1985a; Ma and Ali
2009). On the other hand, POM is composed of plant debris, algae, phytoplank-ton cell, bacteria, and so on (Mostofa et al. 2009a). Humic substances (fulvic and humic acids) of terrestrial origin are the dominant DOM fractions in freshwater and coastal seawater (Mostofa et al. 2009a). On the other hand, autochthonous fulvic acids (or marine humic-like) of algal or phytoplankton and bacterial origin are the key DOM fractions in lakes and oceans (Mostofa et al. 2009a, b; Coble
1996, 2007; Parlanti et al. 2000; Amado et al. 2007; Zhang et al. 2009). In addi-
tion, among the major classes of DOM components there are carbohydrates, pro-teins, amino acids, lipids, phenols, alcohols, organic acids and sterols (Mostofa
et al. 2009a).
DOM can display physical properties such as the absorption of energy from
ultraviolet (UV) and photosynthetically available radiation (PAR) (Kirk 1976; Morris et al. 1995; Siegel and Michaels 1996; Morris and Hargreaves 1997; Tranvik 1998; Bertilsson and Tranvik 2000; Laurion et al. 2000; Markager and
Vincent 2000; Huovinen et al. 2003; Sommaruga and Augustin 2006; Hayakawa
and Sugiyama 2008; Effler et al. 2010), chemical properties such as complex for –
mation with trace metal ions (Mostofa et al. 2009a, 2011; Lead et al. 1999; Wang
and Guo 2000; Koukal et al. 2003; Mylon et al. 2003; Wu et al. 2004; Lamelas and Slaveykova 2007; Lamelas et al. 2009; Fletcher et al. 2010; Reiller and Brevet 2010; Sachs et al. 2010; Da Costa et al. 2011), the ability to maintain acidity and alkalinity (Mostofa et al. 2009a; Oliver et al. 1983; Wigington et al. 1996; Pace and Cole 2002; Hudson et al. 2003; Kopáćek et al. 2003), the occurrence of redox and photo-Fenton reactions (V oelker and Sulzberger 1996; V oelker et al. 1997,
2000; Kwan and V oelker 2002; Jeong and Yoon 2004; Wu et al. 2005; Vione et al. 2006; Nakatani et al. 2007), as well as the ability to control the cycling of nutri-
ents such as NH
4+, NO 3+, and PO 43− in natural waters (Bronk 2002; Zhang et al.
2004, 2008; Kim et al. 2006; Vähätalo and Järvinen 2007; Li et al. 2008).
DOM can photolytically generate strong oxidants such as superoxide radi-
cal (O 2•−), hydrogen peroxide (H 2O2), and hydroxyl radical (HO•), which also
play a role in its photoinduced decomposition in natural waters (Mostofa and Sakugawa 2009; Vione et al. 2006, 2010; Zellner et al. 1990; Zepp et al. 1992;
Moran et al. 2000; Farias et al. 2007; Mostofa et al. 2007a; Minakata et al. 2009). Correspondingly, DOM can undergo photoinduced and microbial degradation processes, which can produce a number of degradation products such as dis-solved inorganic carbon (DIC), CO
2, CH 4, CO, low molecular weight (LMW)
DOM, organic acids. These compounds are very important in the aquatic envi-ronments (Jones and Amador 1993; Miller and Zepp 1995; Lovley and Chapelle 1995; Lovley et al. 1996; Moran and Zepp 1997; Miller 1998; Conrad 1999;
Johannessen and Miller 2001; Ma and Green 2004; Xie et al. 2004; Johannessen et al. 2007; Yoshioka et al. 2007; Brandt et al. 2009; Rutledge et al. 2010; Omar

3 Dissolved Organic Matter in Natural Waters
et al. 2010; Ballaré et al. 2011; Zepp et al. 2011). DOM with its degradation prod-
ucts can extensively influence photosynthesis, thereby playing a key role in global
carbon cycle processes (Mostofa et al. 2009a; Mostofa and Sakugawa 2009; Ma
and Green 2004; Johannessen et al. 2007; Palenik and Morel 1988; Fujiwara et al. 1993; Komissarov 1994, 1995, 2003; Miller and Moran 1997; Meriläinen et al.
2001; Malkin et al. 2008). DOM also plays important roles in regulating drinking water quality, complexing behavior with metal ions, water photochemistry, biolog-ical activity, photosyhthesis, and finally global warming.
This chapter will provide an overview on the origin of DOM, its contents and
sources in natural waters, the contribution of organic substances to DOM, the bio-geochemical functions of DOM, its physical and chemical properties, as well as its molecular size distribution. It comprehensively discusses the controlling fac-tors and their effects on the distribution of DOM in natural waters, the emerging contaminants and their sources, transportation and impacts, as well as methodol-ogies and techniques for the detection of pharmaceuticals in fish tissue. Finally, it is discussed how DOM acts as energy source for living organisms and aquatic ecosystems.
2 What is Dissolved Organic Matter?
DOM is conventionally defined as any organic material that passes through a given filter (0.1–0.7 μm). The organic material that is retained on the filter is termed ‘particulate organic matter (POM) (Mostofa et al. 2009a). The permeate from ultrafiltration (<10 kiloDaltons or kDa) is often defined as the truly dis-solved organic carbon fraction and the filter-passing fraction between >10 kDa and <0.4 or 0.7 μm as the total dissolved organic carbon fraction in aqueous solution. Colloids are operationally defined as particles between 1 nm and 1 μm in size, and the ‘dissolved’ fraction can include a subset of the colloidal materials (Sharp 1973; V old and V old 1983; Koike et al. 1990; Benner et al. 1992; Buesseler et al. 1996;
Wells 2002). These types of colloidal particles are not entirely retained by filters
with pore sizes between 0.2 and 0.7 μm. DOM can be in the size range of tens to hundreds of nm when they are associated with other colloidal materials in water (Lead and Wilkinson 2006). It has been shown that colloids make up a significant
fraction, approximately 10–40 %, of the marine DOM pool.
DOM in natural waters is composed of a heterogeneous mixture of numerous
allochthonous and autochthonous organic compounds containing low molecu-lar weight substances (e.g. organic acids) and macromolecules such as fulvic and humic acids (humic substances), with molecular weight ranging from less than 100 to over 300,000 Daltons (Thurman 1985a, 1986; Ma and Ali 2009; Rashid
and King 1969; MacFarlane 1978; Hayase and Tsubota 1983; Amy et al. 1987; Wagoner et al. 1997; Jerry and Jean-Philippe 2003; Xiao and Wu 2011). DOM found in natural ground and surface waters are also referred as natural organic matter (NOM). The most common organic substances are humic substances

4 K. M. G. Mostofa et al.
(fulvic and humic acids) of terrestrial origin, autochthonous fulvic acids of phy-
toplankton or algal origin, carbohydrates, sugars, amino acids, proteins, lipids, organic acids, phenols, alcohols, acetylated amino sugars, and so on. On the other hand, POM includes plant debris, detritus, living organisms, bacteria, algae, phy-toplankton, corals, coral reefs, and so on. DOM is considered as the larger pool of organic matter in a variety of waters, which can include more than 90 % of total organic matter (Thurman 1986; Kececioglu et al. 1997).
2.1 Biogeochemical Functions of OM (DOM and POM)
DOM of both allochthonous and autochthonous origin can play multiple functions in photoinduced, chemical, microbial and geochemical processes in natural waters. They can be classified as follows:
(1) Photoinduced functions of DOM. Irradiated DOM can produce H
2O2
(Mostofa and Sakugawa 2009), which in turn can produce the strong oxidiz-
ing agent hydroxyl radical (HO•), either directly by photoinduced dissocia-
tion (H 2O2 + hv → HO•) or by the photo-Fenton reaction. These processes
are involved in the photoinduced degradation of organic compounds (Vione et al. 2006, 2010; Zellner et al. 1990; Zepp et al. 1992; Farias et al. 2007).
DOM undergoes rapid photoinduced decomposition by natural sunlight, and this process is less efficient in waters with high contents of DOM and more efficient with high DOM concentrations (Moran et al. 2000; Ma and Green
2004; Vähätalo et al. 2000; Mostofa et al. 2007b; Vione et al. 2009). DOM
can thus control redox and photo-Fenton reactions in natural waters (V oelker and Sulzberger 1996; V oelker et al. 1997, 2000; Kwan and V oelker 2002;
Jeong and Yoon 2004; Wu et al. 2005; Vione et al. 2006; Nakatani et al. 2007).
The biogeochemical functions of H
2O2 and HO• are discussed in details in
chapters “Photoinduced and Microbial Generation of Hydrogen Peroxide and Organic Peroxides in Natural Waters”, “Photoinduced Generation of Hydroxyl Radical in Natural Waters”.
(2) Microbial functions of OM (DOM and POM). DOM and POM are decom-posed biologically by microorganisms in natural waters (Moran et al. 2000; Mostofa et al. 2007a; Ma and Green 2004; Lovley and Chapelle 1995; Hopkinson et al. 2002; Coble 2007; Koschorreck et al. 2008; Lønborg et al. 2009a, b; Lønborg and Søndergaard 2009). This process can produce new
autochthonous DOM or nutrients in water (Mostofa et al. 2009b; Zhang et al. 2009; Kim et al. 2006; Weiss et al. 1991; Harvey et al. 1995; Yamashita
and Jaffé 2008; Fu et al. 2010; Li et al. 2011), so that DOM is responsible for the maintenance of the microbial loop in natural waters (utilization of DOC by bacteria, consumption and decomposition of bacteria by protozo-ans and release of dissolved organic compounds and CO
2) (Sherr and Sherr
1989; Carrick et al. 1991; Jones 1992; Tranvik 1992). Bioavailable carbon

5 Dissolved Organic Matter in Natural Waters
substrates produced from DOM and OM either photolytically or biologi-
cally can enhance biological productivity in waters (Mostofa et al. 2009a; Bertilsson and Tranvik 1998, 2000; Vähätalo and Järvinen 2007; Lovley et
al. 1996; Komissarov 2003; Tranvik 1992; Norrman et al. 1995; Wetzel et
al. 1995). Production of nutrients and DIC through photoinduced and micro-
bial degradation of DOM or POM can control the food-chains for micro-organisms (Mostofa et al. 2009a; Miller and Zepp 1995; Ma and Green 2004; Tranvik 1992; Salonen et al. 1992; Kirchman et al. 1995; Wheeler
et al. 1997; Guildford and Hecky 2000; Rosenstock et al. 2005). The bio-
geochemical functions of microbial processes are discussed in details in “Photoinduced and Microbial Degradation of Dissolved Organic Matter in Natural Waters”.
(3) Optical (or physical) functions of DOM: a fraction of DOM is named as either colored and chromophoric dissolved organic matter (CDOM) based on the absorption of ultraviolet (UV) and photosynthetically available radiation (PAR), or fluorescent DOM (FDOM) based on the emission of fluorescence photons after radiation absorption. DOM generally controls the downward irradiance flux through the water column of UV-B (280–320 nm), UV-A (320–400 nm), total UV (280–400 nm) as well as photosynthetically avail-able radiation (PAR, 400–700 nm) (Kirk 1976; Morris et al. 1995; Siegel and Michaels 1996; Morris and Hargreaves 1997; Tranvik 1998; Bertilsson and
Tranvik 2000; Laurion et al. 2000; Markager and Vincent 2000; Huovinen
et al. 2003; Sommaruga and Augustin 2006; Hayakawa and Sugiyama 2008; Effler et al. 2010). DOM is responsible for water color, water trans-parency, occurrence of the euphotic zone and thermal stratification in the surface waters of lakes and oceans because it affects (decreases) the pen-etration of solar radiation (Laurion et al. 2000; Effler et al. 2010; Hudson et al. 2003; Eloranta 1978; Jones and Arvola 1984; Howell and Pollock 1986;
Perez-Fuentetaja et al. 1999; Snicins and Gunn 2000; Watts et al. 2001; Mostofa et al. 2005a). Biogeochemical functions of CDOM and FDOM are discussed in detail in the respective chapters (see chapters “Colored and Chromophoric Dissolved Organic Matter (CDOM) in Natural Waters” and “Fluorescent Dissolved Organic Matter in Natural Waters”).
(4) Cycling of nutrients (NH
4+, NO 3+, and PO 43−) by DOM and POM.
Nutrients are produced by degradation of DOM and typically derive from dissolved organic nitrogen (DON) and dissolved organic phosphorus (DOP) in DOM molecular structure (Bronk 2002; Zhang et al. 2004, 2008;
Kim et al. 2006; Vähätalo and Järvinen 2007; Li et al. 2008). Nutrients are mostly released during the photoinduced and microbial respiration (or assimilation) of POM (e.g. algae or phytoplankton biomass), as shown by in situ experiments conducted under light and dark incubations (Kim et al. 2006; Li et al. 2008; Yamashita and Jaffé 2008; Carrillo et al. 2002;
Kopáček et al. 2004; Fu et al. 2005; Mostofa KMG et al. unpublished data). NO
3− and NO 2− can be produced by oxidation of ammonia in nitrification
(NH 4+ + 2O 2 → NO 3− + 2H+ + H 2O) and of DON in lake waters (Ma and

6 K. M. G. Mostofa et al.
Green 2004; Kopáček et al. 2004; Lehmann et al. 2004; Minero et al. 2007).
Nutrients produced by DOM and OM can fuel new primary and second-
ary production in natural waters. Total contents of DOM in lake waters are responsible for variation of the trophic level, due to eutrophication/oligotrophication processes. The latter are a major driver of change for chemical vari-ables such as major ions, nutrients (phosphorus and nitrogen compounds, sil-ica) and the chemical nature of DOM.
(5) DOM can control photosynthesis in natural waters. DOC can limit produc-tivity (Jackson and Hecky 1980; Carpenter et al. 1998) and affect epilim-netic (Hanson et al. 2003) and hypolimnetic respiration (Houser et al. 2003).
Photoinduced and microbial oxidation of DOM is responsible for the simul-taneous generation of H
2O2, CO 2 and DIC (Mostofa and Sakugawa 2009;
Ma and Green 2004; Johannessen et al. 2007; Palenik and Morel 1988; Fujiwara et al. 1993; Miller and Moran 1997; Meriläinen et al. 2001; Malkin et al. 2008). Such compounds could favor the occurrence of photosynthesis in natural waters. Some studies show that H
2O2 could be involved as reac-
tant in photosynthesis (xCO 2 + yH 2O2(H2O) + hυ → C x(H2O)y + O 2 + ene
rgy; and 2H 2O2 + hυ → 2H 2O + O 2) (Mostofa et al. 2009a; Komissarov
1994, 1995, 2003; Miller and Moran 1997). Nutrients (PO 43− and NH 4+)
released by DOM and POM might also favor the occurrence of photosynthe-sis and subsequently enhance the cyanobacterial or algal blooms in natural waters (Zhang et al. 2008, 2009; Kim et al. 2006; Li et al. 2008; Lehmann
et al. 2004; Huszar et al. 2006; Nõges et al. 2008; McCarthy et al. 2009;
Mohlin and Wulff 2009). High chlorophyll a concentrations are often detected
in waters with high contents of DOM, and the reverse happens in low-DOM waters (Meriläinen et al. 2001; Malkin et al. 2008; Fu et al. 2010; Guildford and Hecky 2000; Mostofa et al. 2005a, Mostofa KMG et al., unpublished data; Satoh et al. 2006; Yacobi 2006; Komatsu et al. 2007).
(6) Chemical functions of OM (DOM and POM). DOM and POM are com-posed of various functional groups in their molecular structures, which can form complexes with trace metal ions (M) in aqueous solution via strong π-electron bonding systems (Mostofa et al. 2009a, 2011; Lead et al.
1999; Wang and Guo 2000; Koukal et al. 2003; Mylon et al. 2003; Wu
et al. 2004; Lamelas and Slaveykova 2007; Lamelas et al. 2009; Fletcher et al. 2010; Reiller and Brevet 2010; Sachs et al. 2010; Da Costa et al. 2011).
These studies imply that the M-DOM complexation is important for specia-tion, bioavailability, transport and ultimate fate of trace metal ions in the water environment. The detailed functions of M-DOM complexes are discussed in Complexation of Dissolved Organic Matter With Trace Metal Ions in Natural Waters. DOM can also influence the cycling of aluminum and iron oxides in natural waters (McKnight et al. 1992).
(7) Maintenance of the drinking water quality by DOM and POM in waters (Mostofa et al. 2009a). The production of POM is significantly dependent on the DOM contents in natural waters, and POM can produce new autoch-thonous DOM and nutrients under both irradiation and microbial respiration

7 Dissolved Organic Matter in Natural Waters
or assimilation (Mostofa et al. 2005a, 2009b; Zhang et al. 2009; Kim
et al. 2006; Li et al. 2008; Yamashita and Jaffé 2008; Carrillo et al. 2002;
Kopáč ek et al. 2004; Fu et al. 2005). Simultaneously, DOM can release
nutrients upon exposure to natural sunlight in waters (Bronk 2002; Zhang et al. 2004, 2008; Kim et al. 2006; Vähätalo and Järvinen 2007; Li et al.
2008). Increases in nutrients and autochthonous DOM severely deteriorate the drinking water quality, but DOM can also balance acidity and alkalinity through its photoinduced or microbial decomposition (Mostofa et al. 2009a; Oliver et al. 1983; Wigington et al. 1996; Pace and Cole 2002; Hudson et al. 2003; Kopáć ek et al. 2003).
(8) OM can maintain global carbon cycle processes through production, distribu-tion, transportation and decomposition of carbon compounds in the biosphere (Mostofa et al. 2009a; Brandt et al. 2009; Rutledge et al. 2010; Omar et al. 2010; Ballaré et al. 2011; Zepp et al. 2011; Hedges 1992; Amon and Benner 1994; Ogawa and Tanoue 2003; Freeman et al. 2004; Lavoie et al. 2005; Fenner et al. 2007a, b; Wolf et al. 2007). The photoinduced and microbial
decomposition of DOM and POM yields CO
2, CO, CH 4, DIC (DIC is defined
jointly as dissolved CO 2, H2CO3, HCO 3−, and CO 32−), low molecular weight
DOM and other inorganic ions (Jones and Amador 1993; Miller and Zepp
1995; Lovley and Chapelle 1995; Lovley et al. 1996; Moran and Zepp 1997; Miller 1998; Conrad 1999; Johannessen and Miller 2001; Ma and Green
2004; Xie et al. 2004; Johannessen et al. 2007; Yoshioka et al. 2007; Brandt
et al. 2009; Rutledge et al. 2010; Omar et al. 2010; Ballaré et al. 2011; Zepp et al. 2011). The produced CO
2 and CH 4 increase the atmospheric green
house gases and contribute to the global carbon cycle (Davidson and Janssens 2006; Porcal et al. 2009). Elevated atmospheric CO
2 can enhance DOC sup-
ply, particularly in peat soils. This is attributed to elevated net primary pro-ductivity of plants and increased root exudation of DOC in soil environments, which ultimately leach into the aquatic ecosystem (Freeman et al. 2004; Lavoie et al. 2005; Fenner et al. 2007a, b; Wolf et al. 2007; Kang et al. 2001;
Pastor et al. 2003).
(9) Character and energy functions of OM in the water ecosystem. DOM and POM can provide a major source of energy, in the form of C and N, which are essential to all living organisms in natural waters (Mostofa et al. 2009a; Tranvik 1992; Salonen et al. 1992; Wetzel 1984, 1992). Thermal energy
produced during the photoinduced and microbial degradation of DOM and organic matter, photoinduced redox reactions, microbial loop, as well as photosynthesis are key drivers in aquatic ecosystems (Mostofa et al. 2009a; Komissarov 1994, 1995, 2003; Miller and Moran 1997; Sherr and Sherr 1989;
Carrick et al. 1991; Jones 1992; Tranvik 1992; Salonen et al. 1992; Wetzel
1984, 1992; Hedges et al. 2000). DOM itself can provide energy and matter
for the growth of bacterial films on the surface of drinking-water pipes, a pro-cess that involves also fulvic and humic acids (humic substances) depending on their occurrence in groundwater in developing and developed countries (Mostofa et al. 2009a).

8 K. M. G. Mostofa et al.
3 Origin of DOM in Natural Waters
DOM is generally originated from three major sources in natural waters: (i) DOM
derived from terrestrial soils, termed allochthonous DOM; (ii) DOM derived from in situ production in natural surface waters, termed autochthonous DOM, and (iii) DOM derived from human activities (e.g. industrial synthesis), termed anthropo-genic DOM.
3.1 Origin of Allochthonous DOM in Soil Ecosystems
DOM including fulvic and humic acids (humic substances) originates from the decomposition of vascular plant material, root exudates and animal residues in ter –
restrial soil. Origin of allochthonous DOM from vascular plant materials or partic-ulate detrital pools is significantly varied in different regions (tropical, temperate and boreal), which is regulated by the occurrence of three key factors or functions (Mostofa et al. 2009a; Wetzel 1983, 1990, 1992; Malcolm 1985; Dai et al. 1996;
Nakane et al. 1997; Wershaw 1999; Jaramillo and Dilcher 2000; Kalbitz et al.
2000; Trumbore 2000; Uchida et al. 1998, 2000; Moore et al. 2008; Braakhekke
et al. 2011; Spence et al. 2011; Tu et al. 2011): (i) Physical functions that include temperature and moisture; (ii) Chemical functions that include nutrient avail-ability, amount of available free oxygen and redox activity, and (iii) Microbial processes that include microfloral succession patterns and availability of microor –
ganisms (aerobic or anaerobic).
It is suggested that microorganisms can alter sugars, starch, proteins, cellulose
and other carbon compounds bound to organic matter of plant or animal origin during their metabolic processes. These processes can transform the aromatic and lipid plant components into amphiphilic molecules including humic substances, i.e., molecules that consist of separate hydrophobic (non-polar) and hydrophilic (polar) parts (Wershaw 1999). The non-polar parts of the molecules are composed
of relatively unaltered segments of plant polymers, while the polar parts include carboxylic acid groups (Wershaw 1999). Aerobic microorganisms can decompose
organic matter at a faster rate than anaerobic ones, depending on the availability of free oxygen. Compositional changes of DOM occur with soil depth, leading to a decrease of aromatic compounds and carbohydrates whilst alkyl, methoxy and carbonyl moieties increase with depth (Dai et al. 1996). The increase in alkyl and
carboxylic C with depth are the result of biodegradation of forest litter and oxi-dation of lignin side chains, respectively (Zech et al. 1985; Kogel-Knabner et al. 1988; Kogel-Knabner 1992).
The origin of allochthonous DOM from microbial processes can be judged
from significant variations in respired organic carbon in different soil environ-ments. The mean age of soil respired organic carbon determined using
14C tracer
is lowest (1 year) in tropical forest soils (eastern Amazonia, Brazil), relatively

9 Dissolved Organic Matter in Natural Waters
low (3 years) in temperate forest soils (central Massachusetts, USA), and high-
est (16 years) in boreal forest soils (Manitoba, Canada) (Trumbore 2000). Experimental studies using δ
13C or 14C to track sources and turnover of DOC
indicate that DOM, which is transported over decimetres or metres down into sub-soil, mainly represents highly altered residues of organic matter processing (Schiff et al. 1997; Flessa et al. 2000; Hagedorn et al. 2004; Fröberg et al. 2007). Note
that allochthonous DOM is mostly derived, in zero to a few decimeter depth from the decomposition of plant material by microbial processes in soils and shallow groundwater (Uchida et al. 1998, 2000; Fröberg et al. 2007; IPCC 1996; Buckau et
al. 2000).
DOC leached from soil is partly retained in the vadose zone before reach-
ing aquifers (Siemens and Kaupenjohann 2003; Mikutta et al. 2007; Kalbitz and Kaiser 2008; Scheel et al. 2008). For the range of groundwater recharge of 95–652 mm yr
−1, it is shown that a constant flux of DOC from soil into surface
waters often takes place (Kindler et al. 2011). Therefore, allochthonous DOM is partly discharged through hydrological processes directly into streams or riverbeds or surrounding water bodies, which ultimately flux to lake or oceanic environ-ments as final water reservoir.
3.2 Origin of Autochthonous DOM in Natural Waters
Production of autochthonous DOM is generally observed at the epilimnion (upper water layers) compared to the hypolimnion (deeper layers) during the sum-mer stratification period, particularly in lakes and oceans. A rough estimation by comparing the upper with the deeper layers demonstrates that the contribution of autochthonous DOM is largely varied in lakes and oceans: it reaches 0–55 % in Lake Hongfeng (181–250 μM C at 0–6 m and 161–223 μM C at 22–25 m depth, respectively, during March–September), 3–47 % in Lake Baihua (183–264 μM C at 0–3 m and 157–206 μM C at 14–15 m during March-September), 6–35 % in Lake Baikal (93–142 μM C at 0–100 m and 88–105 μM C at 600–720 m during August–September in 1995, 1998, 1999), 3–82 % in Lake Biwa (93–183 μM C at 2.5–10 m and 78–101 μM C at 70 m during May–September in 1999–2002), 21–49 % in Lake Ashino in Japan (99–111 μM C at 0–10 m and 74–84 μM C at 30–38 m in September 1997), 81–102 % in Lake Ikeda in Japan (101–112 μM C
at 0–10 m and 55–56 μM C at 200–233 m for site I1; at 41 m for site I2 in October 1997), 52 % in Lake Suwa in Japan (216 μM C at 0 m in September and 142 μM C at 0 m in December 1997), 61–81 % in Lake Inawashiro in Japan (42–
47 μM C at 0–10 m and 26 μM C at 70 m), 13–29 % in Lake Fuxian (123–135
μM C at 0–10 m and 95–105 μM C at 50–140 m in June 2001), 19 % in Lake Hovsgol (95 μM C at 0 m and 80 μM C at 50–200 m in July 1999), 0–88 % in Lake Kinneret (270–485 μM C at 0–10 m and 258–368 μM C at 38 m during the summer period in 2004), 17–41 % in Lake Peter (data not shown), 11–29 % (bio-logical production) in Lake Bret, 0–104 % in Middle Atlantic Bight (82–98 μM C

10 K. M. G. Mostofa et al.
at 0 m and 48–90 μM C at 90–2600 m in June 2001), 16–77 % in Western North
Pacific (85–117 μM C at 0 m and 66–73 μM C at 150 m), 0–194 % in Atlantic Ocean (50–97 μM C at <100 m and 33–59 μM C at >1000 m), 0–165 % in Pacific Ocean (40–90 μM C at <100 m and 34–45 μM C at >1000 m), 28–121 % in Indian Ocean and Arabian Sea (55–95 μM C at <100 m and 43 μM C at >1000 m), 0–121 in Antarctic Ocean (38–75 μM C at <100 m and 34–60 μM C at >1000 m), as well as 0–118 % in Arctic Ocean (34–107 μM C at <100 m and 49–54 μM C at >1000 m) (Mostofa et al. 2005a, 2009a; Fu et al. 2010; Ogawa
and Tanoue 2003; Ogawa and Ogura 1992; Wilkinson et al. 1997; Mitra et al.
2000; Yoshioka et al. 2002a; Hayakawa et al. 2003, 2004; Annual Report 2004;
Bade 2004; Sugiyama et al. 2004).
The contribution of extracellular release of photosynthetically-derived DOM
varies from 5 to 70 % in natural waters (Lancelot 1979; Fogg 1983; Connolly
et al. 1992). The autochthonous production is significantly higher in oceans with a high water temperature (WT) than in those with a low water temperature, par –
ticularly in the Arctic Ocean. The key contributors to autochthonous DOM in natural waters as well as in sediment pore waters are considered to be phyto-plankton or algal biomass, bacteria, coral, coral reef, submerged aquatic vegeta-tion, krill (shrimp-like marine crustaceans), seagrass, and marsh- and mangrove forest (Mostofa et al. 2009a, b; Zhang et al. 2009; Li et al. 2011; McKnight et al.
1991, 1993, 1994, 2001; Tanoue et al. 1995, 1996; Fukuda et al. 1998; Nelson et
al. 1998, 2004; Tanoue 2000; Kahru and Mitchell 2001; Ogawa et al. 2001; Hata
et al. 2002; Rochelle-Newall and Fisher 2002a, b; Burdige et al. 2004; Cammack
et al. 2004; Steinberg et al. 2004; Wild et al. 2004; Yamashita and Tanoue 2004;
Biers et al. 2007; Chen et al. 2007; Vantrepotte et al. 2007; Wada et al. 2007; Wang et al. 2007; Hanamachi et al. 2008; Henderson et al. 2008; Tanaka et al. 2008; Tzortziou et al. 2008; Ortega-Retuerta et al. 2009; Tranvik et al. 2009). These studies demonstrate that autochthonous DOM is produced from POM by several processes such as photoinduced and microbial respiration (or assimilation), zooplankton grazing, bacterial release and uptake, viral interactions, and complex microbial processes in sediment pore waters.
3.2.1 Respiration or Assimilation of Algae or Phytoplankton
Species and Bacteria
Algae or phytoplankton biomass and bacteria can release new DOM in natu-ral waters by two key processes: first, photoinduced respiration or assimilation of algae or phytoplankton biomass and bacteria, which can produce new DOM (Mostofa et al. 2005a, 2009b, 2011; Rochelle-Newall and Fisher 2002a; Varela
et al. 2003; Aoki et al. 2008; Biddanda and Benner 1997; Hulatt et al. 2009). Second, microbial respiration or assimilation of algae or phytoplankton and bac-teria, which can release the new DOM in natural waters (Mostofa et al. 2009a, b,
2011; Parlanti et al. 2000; Zhang et al. 2009; Fu et al. 2010; McKnight et al. 1991,
1994, 2001; Nelson et al. 2004; Rochelle-Newall and Fisher 2002a; Cammack

11 Dissolved Organic Matter in Natural Waters
et al. 2004; Yamashita and Tanoue 2004, 2008; Wada et al. 2007; Hanamachi et al.
2008; Ortega-Retuerta et al. 2009; Aoki et al. 2008; Biddanda and Benner 1997;
Hulatt et al. 2009; Bertilsson and Jones 2003; Chen and Gardner 2004; Stedmon
and Markager 2005a; Stedmon et al. 2007a, b; Wetz and Wheeler 2007; Zhao et al.
2009).
Re-suspension of algae or phytoplankton in ultrapure water (Milli-Q), arti-
ficial sea water and natural waters can release new organic compounds, either under irradiation or under dark incubation. These organic substances, produced either under irradiation (Fig. 1a) or in the dark (Fig. 1b, c) show fluorescence (excitation-emission matrix, EEM) properties. The EEM spectra of autochtho-nous DOM (Fig. 1a, b) are roughly similar to those of allochthonous fulvic acid and show two fluorescence peaks at peak C- and A-regions (Fig. 1d). In contrast, they are different from allochthonous humic acids that show more than two peaks at peak C-region (Fig. 1f). Based on the similarities of the EEM spectra, the key
component of autochthonous fluorescent DOM is defined as “autochthonous ful-vic acid (C-like)” of algal or phytoplankton origin. The other component (Fig. 1c) is defined as “autochthonous fulvic acid (M-like)” of algal or phytoplankton ori-gin, based on the similarities with the marine humic-like component (Coble 1996, 2007). Identification of autochthonous DOM of algal or phytoplankton origin is
Peak A Peak C
Peak M(b) (a)
Ex wavelength (nm)Em wavelength (nm)(e)(d)Peak A Peak C
Peak APeak C(c)
Peak M
Peak A Peak C
Fig. 1 Comparison of the fluorescent components of autochthonous fulvic acid (C-like) pro-
duced under microbial respiration of lake algae (a), autochthonous fulvic acid (C-like) under
photorespiration or assimilation of algal biomass (b) and autochthonous fulvic acid (M-like) under microbial respiration of algae (c) with aqueous samples of standard Suwannee River Ful-vic Acid (d) and Suwannee River Humic Acid (e) identified using PARAFAC modeling on the EEM spectra of their respective samples. Data source Mostofa KMG et al., (unpublished data)

12 K. M. G. Mostofa et al.
discussed extensively in the FDOM chapter (see chapter “Fluorescent Dissolved
Organic Matter in Natural Waters”). Note that “autochthonous fulvic acids” of algal or phytoplankton origin are newly termed in this study for mostly two rea-sons: first, to distinguish and generalize between all freshwaters and marine waters; second, because of the confusion in different studies that use several names such as marine humic-like (Coble 1996, 2007), sedimentary fulvic acids
(Hayase and Tsubota 1983), microbially derived fulvic acids or marine fulvic
acids (McKnight et al. 1991, 1994; Harvey and Boran 1985; Meyers-Schulte and
Hedges 1986).
DOM is produced significantly by eleven species of intertidal and sub-tidal
macroalgae when they are illuminated, providing evidence for a light-driven exu-dation mechanism (Hulatt et al. 2009). The contribution of the released DOC has been detected as 6.4 and 17.3 % of the total organic carbon in cultures of Chlorella vulgaris and Dunaliella tertiolecta, respectively, upon light exposure (Hulatt and Thomas 2010). DOM can support a significant growth of bacterial bio-
mass, representing a further loss of algal assimilated carbon in water (Hulatt and Thomas 2010). Dissolved combined amino acids, middle-reach peaks of particu-
late amino acids and non-protein amino acids are often decreased in downstream rivers, which is likely the result of photoinduced degradation of DOM and algae (Duan and Bianchi 2007).
On the other hand, the key processes of autochthonous DOM release by micro-
bial respiration of algae or phytoplankton biomass in waters are presumably the extracellular release by living cells, cell death and lysis, or herbivore grazing that may occur in the deeper waters of rivers, lakes and oceans (Mostofa et al. 2009a; Tanoue 2000; Tranvik et al. 2009; Hulatt et al. 2009). In fact, bacteria play
a specific role in subsequent processing of the DOM released by algae in natu-ral water (Nelson et al. 1998, 2004; Rochelle-Newall and Fisher 2002a; Cammack
et al. 2004; Biers et al. 2007; Ortega-Retuerta et al. 2009). Cultivation of three
kinds of phytoplankton (green algae Microcystis aeruginosa and Staurastrum dor –
cidentiferum and dark-brown whip-hair algae Cryptomonas ovata collected from lake waters) shows that fulvic acid-like and protein-like fluorescent components are released when they are cultivated under a 12:12 h light/dark cycle in a MA medium and an improved VT medium at 20 °C (Aoki et al. 2008). This study implies that the increase of the refractory organic matter in lake waters may be attributed to a change of the predominant phytoplankton. Similarly, cultivation of three kinds of phytoplankton (Prorocentrum donghaiense, Heterosigma akashiwo and Skeletonema costatum collected from sea water) can produce visible humic-
like (C-like and M-like) and protein-like or tyrosine-like components in waters (Zhao et al. 2006a, 2009).
Releases of DOM by eleven species of intertidal and sub-tidal macroal-
gae in the dark account for 63.7 % of that in the light in the UV-B band (Hulatt
et al. 2009). Some brown algae can produce considerably less DOM (e.g. Pelvetia
canaliculata), which are more comparable to the green and red species (Hulatt et al. 2009). It is shown that thin, subsurface DOM maxima are observed below the plume during the highly stratified summer period but are absent in the spring,

13 Dissolved Organic Matter in Natural Waters
which is the strong evidence of significant in situ biological production of CDOM
in natural waters (Chen and Gardner 2004).
Incubation of coastal seawater in the presence of model (DON: amino sug-
ars and amino acids) and DIN compounds shows that net biological DOM for –
mation occurs upon addition of amino sugars (formation of fluorescent, mostly labile DOM) and tryptophan (formation of non-fluorescent, refractory DOM) (Biers et al. 2007). Similarly, natural assemblages of marine bacteria can rap-idly produce refractory material (in <48 h) utilizing labile compounds (glucose, glutamate), as observed in a laboratory experiment (Ogawa et al. 2001). On the
other hand, photoinduced formation of DOM is only detected when tryptophan is added to the water (Biers et al. 2007). This CDOM is highly fluorescent, with excitation-emission matrices (EEMs) resembling those of terrestrial, humic-like fluorophores (Biers et al. 2007). The bulk particulate organic carbon (POC) dur –
ing the decomposition process of freshwater or marine algae and phytoplank-ton is significantly decreased during the first few days. It subsequently remains almost constant (Zhang et al. 2009; Hanamachi et al. 2008; Matsunaga 1981; Fukami et al. 1985; Osinga et al. 1997; Fujii et al. 2002). The carbohydrate con-tents of both the particulate and dissolved pools are increased during the phy-toplankton growth cycle, accounting for 18–45 % and 26–80 % of total organic carbon (TOC), respectively, in natural waters (Biddanda and Benner 1997). Photoreactions driven by UV-B can reduce the microbial availability of certain organic substrates such as peptone and algal exudates (Morris and Hargreaves 1997; Thomas and Lara 1995; Naganuma et al. 1996). This phenomenon can be caused by light-induced cross-linking between DOM and algal exudates (Morris and Hargreaves 1997).
LMW organic acids are presumably formed by four major processes (Lovley
et al. 1996; Xiao and Wu 2011; Wetzel et al. 1995; Smith and Oremland 1983; Kieber et al. 1990; Corin et al. 1996; Janczarek et al. 1997; Evans 1998; Bertilsson et al. 1999; Tedetti et al. 2006; Lu et al. 2007; Xiao et al. 2009, 2011): first, pho-
toinduced decomposition of allochthonous and autochthonous DOM in surface waters; second, photoinduced and microbial respiration or assimilation of algae or phytoplankton biomass in natural waters; third, conversion of anaerobic organic substances (carbohydrates, fats, proteins, etc.) into CH
4 and CO 2 in pore waters or
soil ecosystems; and fourth, root exudations of plants or plant–microbe associa-tions (e.g. Rhizobium symbiosis with leguminous roots).
A number of factors can influence the DOC release by algae or phytoplank-
ton and bacteria in waters, which can be distinguished as: (i) occurrence of the phytoplankton species and their contents; (ii) water quality; (iii) presence of nutrients; (iv) effect of UV and PAR; (v) water temperature; (vi) occurrence of microbes; (vii) metabolic abilities or inabilities and so on (Norrman et al. 1995; Mostofa KMG et al., unpublished data; Lancelot 1979; Fogg 1983; Nelson et al.
1998, 2004; Rochelle-Newall and Fisher 2002a, b; Cammack et al. 2004; Biers
et al. 2007; Ortega-Retuerta et al. 2009; Hulatt et al. 2009; Zhao et al. 2006a, 2009; Williams 1990, 1995; Obernosterer and Herndl 1995; Anderson and
Williams 1998; McCallister et al. 2004).

14 K. M. G. Mostofa et al.
3.2.2 Photosynthesis
Photosynthesis is the key process for the formation of organic carbon or OM
(e.g. algae or cyanobacteria, phytoplankton, etc.) through light-stimulated inor –
ganic carbon acquisition in surface waters (Mostofa et al. 2009a; Komissarov
1994, 1995, 2003; Li et al. 2011; Li 1994; Zubkov and Tarran 2008; Beardall
et al. 2009a, b; Wu and Gao 2009; Liu et al. 2010). Photosynthetic organisms are
then able to produce autochthonous DOM via photoinduced respiration (or pho-toinduced assimilation) and microbial respiration or assimilation in natural waters (Mostofa et al. 2009b; Zhang et al. 2009; Conrad 1999; Weiss et al. 1991; Harvey et al. 1995; Fu et al. 2010; Thomas and Lara 1995; Druon et al. 2010; Yamashita
et al. 2008). A new hypothesis on photosynthesis also considers that H
2O2 might
be involved in the occurrence of oxygenic photosynthesis in both higher plants (Komissarov 1994, 1995, 2003; Miller and Moran 1997) and natural water organ-
isms (Mostofa et al. 2009a, b). Occurrence of photosynthesis in natural waters
includes two facts: the first is the generation of numerous chemical species from DOM, which may proceed as follows: (i) photoinduced degradation of DOM can produce many photoproducts, such as H
2O2, CO 2, DIC, CO, LMW DOM, and
so on in upper surface waters (Mostofa and Sakugawa 2009; Miller and Zepp 1995; Miller 1998; Johannessen and Miller 2001; Ma and Green 2004; Xie et al.
2004; Johannessen et al. 2007; Salonen and Vähätalo 1994; Amon and Benner 1996; Granéli et al. 1996; Remington et al. 2011; Zepp et al. 1998; Cai et al. 1999; Gennings et al. 2001; Clark et al. 2004; Fichot and Miller 2010; White et al. 2010; Cai 2011); (ii) microbial degradation of DOM including DON and DOP can pro-
duce compounds such as H
2O2, CO 2, DIC, PO 43−, NH 4+, CH 4, LMW DOM and
so on (Mostofa and Sakugawa 2009; Zhang et al. 2004; Vähätalo and Järvinen 2007; Lovley et al. 1996; Ma and Green 2004; Palenik and Morel 1988; Li et al. 2011; Zinder 1990; Kotsyurbenko et al. 2001; Zagarese et al. 2001; Semiletov
et al. 2007). Many of these compounds can favor the occurrence of photosynthe-sis either directly or indirectly and lead to fixation of organic carbon or OM from inorganic carbon in surface waters (Mostofa et al. 2009a; Komissarov 1994, 1995,
2003; Miller and Moran 1997; Li et al. 2011; Ortega-Retuerta et al. 2009; Li 1994;
Zubkov and Tarran 2008; Beardall et al. 2009a, b; Wu and Gao 2009; Liu et al.
2010).
A general scheme for the photoinduced (Eq. 3.1) and microbial or biological
(Eq. 3.2) degradation of DOM can be expressed as follows (Mostofa et al. 2009a, b):
The second fact is that H
2O2 and CO 2, produced by either photoinduced or
microbial degradation of DOM and POM can take part to photosynthesis, to form new OM or carbohydrate-type compounds (Mostofa et al. 2009a, b):(3.1)
DOM +hυ→H2O2+CO2+DIC+CO+LMW DOM
(3.2)DOM +microbes →CO2+DIC+PO43−+NH 4++CH4+LMW DOM

15 Dissolved Organic Matter in Natural Waters
where C x(H2O)y represents a generic carbohydrate (Eq. 3.3). According to this
hypothesis, H 2O2 acts together with carbon dioxide (CO 2) to form carbohydrates
and oxygen (Eq. 3.3). The formation of oxygen occurs via H 2O2 disproportiona-
tion (Eq. 3.4) that is a common conversion reaction of H 2O2 in water ecosystems
and the atmosphere (see the photosynthesis chapter for detailed description for
these reactions) (Liang et al. 2006; Buick 2008). In Eq. (3.3), E (±) is the energy produced during photosynthesis.
Currently, model results imply that the progressive release of DON in
the ocean’s upper layer during summer increases the primary production by 30–300 %. This will in turn enhance the DOC production mostly from phyto-plankton exudation in the upper layer and the solubilization of POM deeper in the water column (Druon et al. 2010). Experimental studies observe that both the quantity and the spectral quality of DOM produced by bacteria can be influenced by the presence of photoproducts in aqueous media (Ortega-Retuerta et al. 2009). Photosynthetically produced POM (algae or phytoplankton) and their photo- and microbial respirations are significantly influenced by several key factors, such as chemical nature and contents of DOM (Jones 1992; Hessen 1985; Tranvik and Hafle 1987; Tranvik 1989); high precipitation (Freeman et al. 2001a; Tranvik and
Jasson 2002; Hejzlar et al. 2003; Zhang et al. 2010); land use changes that cause
high transport of DOC from catchments to adjacent surface waters (Worrall et al. 2004a; Raymond and Oh 2007); nitrogen deposition (Pregitzer et al. 2004; Findlay
2005); sulfate (SO
42−) deposition (Zhang et al. 2010; Evans et al. 2006; Monteith
et al. 2007); droughts and alteration of hydrologic pathways (Hongve et al. 2004; Worrall and Burt 2008); changes in total solar UV radiation or an increase in tem-perature due to global warming (Freeman et al. 2001a; Zhang et al. 2010; Sinha
et al. 2001; Sobek et al. 2007; Rastogi et al. 2010).
Finally, H
2O2 can react with CO 2 under abiotic conditions to produce vari-
ous organic substances (CH 2O, HCOOH, CH 3OH, CH 4, C6H12O6; Eqs. 3.5–3.9,
respectively) in aqueous solution (Lobanov et al. 2004). The reactions between H
2O2 and CO 2 as well as their thermodynamic parameters such as enthalphy
changes (ΔH0) and the Gibbs free energy changes (ΔG0) are as follows (Lobanov
et al. 2004):(3.3) xCO2(H2O)+yH2O2(H2O)+hυ→Cx(H2O)y+O2+E(±)
(3.4) 2H2O2+hυ→2H2O+O2
(3.5) H2O2+CO2→CH2O+3/2O 2
∆H0=465kJ , ∆G0=402kJ
(3.6) H2O2+CO2→HCOOH +O2
∆H0=172kJ , ∆G0=166kJ

16 K. M. G. Mostofa et al.
3.3 DOM Derived from Anthropogenic and Human Activities
Organic pollutants derived from sewerage and from domestic, agricultural and
industrial effluents significantly contribute to increase the concentration levels of DOM in natural waters (Fu et al. 2010; McCalley et al. 1981; Silberhorn et al. 1990; Kramer et al. 1996; Mudge and Bebianno 1997; Manoli and Samara 1999; Abril et al. 2002; Newton et al. 2003; Mostofa et al. 2005b, 2010; Richardson
2003, 2007; Mottaleb et al. 2005, 2009; Mudge and Duce 2005; Richardson
and Ternes 2005, 2011; Buser et al. 2006; Field et al. 2006; Lishman et al. 2006;
Rudel et al. 2006; Xia et al. 2006; Brown et al. 2007; Schmid et al. 2007; Farré
et al. 2008; Kinney et al. 2008; Guo et al. 2009; Ramirez et al. 2009; Citulski and Farahbakhsh 2010; Kumar and Xagoraraki 2010; Pal et al. 2010; Yoon et al. 2010;
Kleywegt et al. 2011; Yu et al. 2011). The organic matter pollution is an important problem in both developed and developing countries through input of untreated sewerage and industrial effluents into natural waters. However, its impacts may be much worse in developing countries due to the lack of sewerage treatment and of industrial effluent treatment plants. The occurrence of DOM derived from anthro-pogenic and human activities is gradually increasing because of the increasing dif-fusion of domestic, agricultural and industrial activities. Some components of sewerage-impacted DOM are made up of detergents or fluorescent whitening agents (FWAs), including mostly diaminostilbene type (DAS1) and distyryl biphe-nyl (DSBP), protein-like components, sterols, and unknown organics (McCalley
et al. 1981; Mudge and Bebianno 1997; Mostofa et al. 2005b, 2010; Mudge and
Duce 2005). The organic components originating from agricultural wastes are pesti-
cides, herbicides, dichlorodiphenyltrichloroethane (DDT) and their degradation prod-ucts (Richardson 2007; Guo et al. 2009; Derbalah et al. 2003; Medana et al. 2005).
Recent studies show that emerging organic contaminants such as pharmaceuti-
cals and personal care products (PPCPs) are a ubiquitous class of organic chemi-cals of considerable concern for natural waters, and will be discussed in details later. Wastewater-derived organic compounds can produce three major types of toxic byproducts such as trihalomethanes (THMs), N-nitrosodimethylamine (NDMA) and organic chloramines. These compounds may be formed either upon (3.7)
2H2O2+CO2→CH3OH + 5/2O 2
∆H0=530 kJ , ∆G0=464 kJ
(3.8) 2H2O2+CO2→CH4+ 3O 2
∆H0=694 kJ , ∆G0=580 kJ
(3.9) H2O2+CO2→1/6C 6H12O6+ 3/2O 2
∆H0=426 kJ

17 Dissolved Organic Matter in Natural Waters
chlorination or in conventional and advanced wastewater treatment plants (Scully
et al. 1988; Jensen and Helz 1998; Jameel and Helz 1999; Mitch et al. 2003).
4 Contribution of Organic Substances to DOM
in Natural Water
The contributions of major organic substances in streams and rivers to the total
DOM pool are 20–85 % of humic substances, of which 15–80 % fulvic acid and 5–29 % humic acid (the ratio of fulvic acid to humic acid is 9:1 for lower stream DOC and it decreases to 4:1 or less for higher stream DOC), 10–30 % of carbo-hydrates, 2–48 % of dissolved amino acids, organic acids or hydrophilic acids (9–25 %), autochthonous fulvic acids of phytoplankton or algal origin (or marine humic-like: see Sect. 3.2 and also FDOM chapter for detailed description), organic
acids, organic peroxides (ROOHs), sterols; organic contaminants of anthropo-genic origin and so on (Mostofa et al. 2009a; Malcolm 1985, 1990; Bertilsson
et al. 1999; Lu et al. 2007; Wetzel and Manny 1972; Meybeck 1982; Meyer and Tate 1983; Ittekkot et al. 1985; Thurman 1985b; Meyer 1986; Tipping et al. 1988;
Lewis and Saunders 1989; Peuravuori 1992; Hedges et al. 1994; Eatherall 1996; V olk et al. 1997; Kusel and Drake 1999; Peuravuori and Pihlaja 1999; Alberts and Takács 1999; Ma et al. 2001; Raymond and Bauer 2001a; van Hees et al. 2002;
Nagai et al. 2005; Mostofa 2005; Guéguen et al. 2006). Hydrophilic acids gener –
ally include amino acids, proteins, carbohydrates and free sugars. The contribu-tion of humic substances (hydrophobic acids) in groundwater is approximately 12–98 % (1–80 % of fulvic acid and 2–97 % of humic acid), and the contribu-tion of hydrophilic fractions is 1–82 % (Buckau et al. 2000; Bertilsson et al. 1999; Peuravuori and Pihlaja 1999; Leenheer et al. 1974; Thurman 1985c; Ford and
Naiman 1989; Schiff et al. 1990; Wassenaar et al. 1990; Malcolm 1991; Grǿn et al. 1996; Christensen et al. 1998; McIntyre et al. 2005; Mladenov et al. 2008). These studies observe high variation in the contribution of humic substances from stream (source) to the end of river mouths. The main reasons are the mixing up of vari-ous sources of water in the downstream locations as well as the photoinduced and microbial changes during transportation.
In lakes the contributions of humic substances (fulvic and humic acids) account
for 14–90 % of total DOM (14–70 % of fulvic acid and 0–22 % of humic acid); the DOM pool is also made up of ~12–60 % of autochthonous fulvic acids (see FDOM chapter for detailed description) of algal or phytoplankton origin; of car –
bohydrates for 1–65 %; of amino acids, proteins and organic acids that together account for 10–33 % of total DOM; of organic acids (2.5–7.5 %, but 0–11 % in pore water); sterols; algal toxins, organic contaminants of anthropogenic origin and so on (Mostofa et al. 2009a, b; Parlanti et al. 2000; Xiao and Wu
2011; Wilkinson et al. 1997; McKnight et al. 1991, 1994, 1997; Xiao et al. 2009,
2011; Thurman 1985b; Peuravuori 1992; Peuravuori and Pihlaja 1999; Ma et al.
2001; Nagai et al. 2005; Schiff et al. 1990; Steinberg and Muenster 1985; Hama

18 K. M. G. Mostofa et al.
and Handa 1987; Baron et al. 1991; Søndergaard and Middelboe 1995; Reitner et
al. 1997; Malcolm and MacCarthy 1992; Imai et al. 1998; Rosenstock and Simon
2001; Frimmel 2004; Hayakawa 2004; Sugiyama et al. 2005). Biomolecules (e.g.
carbohydrates and proteins) as well as organic acids account for approximately 70 % of high molecular weight (HMW) DOM, and only for approximately 2 % of (LMW) DOM in lake water (Hama and Handa 1992). These studies also show that allochthonous fulvic acids in lakes are largely varied during the summer and winter season, with winter maxima and summer minima. Their total content is also low in algal-dominated lakes.
The percentages of major organic substances in bulk DOM in shelf, coastal
and open ocean are: 1–75 % of allochthonous fulvic acids of terrestrial origin;
5–10 % of autochthonous fulvic acids (or marine humic-like: see Sect. 3.2 and also FDOM chapter for detailed description) of algal or phytoplankton origin; 10–80 % of carbohydrates (~25 % in deeper layers); 10–28 % of amino acids, proteins and lipids taken together (amino acids alone account for 7 %); organic acids; organic peroxides (ROOH); sterols; algal toxins, and so on (Mostofa et al. 2009a, b; Coble 1996, 2007; Zhang et al. 2009; Bronk 2002; Ogawa and Tanoue
2003; Ogawa et al. 2001; Biddanda and Benner 1997; Harvey and Boran 1985; Meyers-Schulte and Hedges 1986; Druon et al. 2010; Richardson 2007; Thurman 1985b; Alberts and Takács 1999; Ma et al. 2001; Beck et al. 1974; Stuermer and
Harvey 1977; Gagosian and Stuermer 1977; Burney et al. 1982; Thurman and
Malcolm 1983; Romankevich 1984; Williams and Druffel 1987; Moran et al.
1991; Moran and Hodson 1994; Pakulski and Benner 1994; McCarthy et al. 1996; Opsahl and Benner 1997; Gašparovic et al. 1998; Kirchman et al. 2001; Aluwihare et al. 2002; Benner and Kaiser 2003; Yamashita and Tanoue 2003). The con-
tributions of allochthonous humic substances in shelf seawater are 11–75 %, of which around 38 % of marsh origin and 62 % of river origin (Moran and Hodson 1994). Carbohydrates can comprise 10–70 % of the organic matter in the plankton
cell (Romankevich 1984) and are presumably released directly to the water col-umn by algae or phytoplankton under photo- and microbial respiration (Mostofa et al. 2009b; Zhang et al. 2009; Hellebust 1965; Ittekkot et al. 1981; Mopper et al. 1995; Cowie and Hedges 1994, 1996). Carbohydrates (originally polysaccha-
rides) make up approximately 15–60 % of marine HMW DOM (Druon et al. 2010; Burney et al. 1982; Romankevich 1984; Pakulski and Benner 1994; McCarthy et
al. 1996). Carbohydrates also account for ~5–20 % of particulate material in sea-
water (Pakulski and Benner 1994; Tanoue and Handa 1987; Hernes et al. 1996; Panagiotopoulos et al. 2002). Autochthonously produced carbohydrates, proteins and lipids are vital biochemical organic groups that together constitute approxi-mately 10–80 % of organic carbon and 15–50 % of the nitrogen assimilated dur –
ing photosynthesis by phytoplankton in natural waters (Sundh 1992; Bronk et al. 1994; Braven et al. 1995; Malinsky-Rushansky and Legrand 1996; Wakeham et al. 1997; Slawyk et al. 1998).
The main organic substances in rainwater are hydrophobic DOM (major frac-
tion; ~<50 %), including allochthonous humic substances (fulvic and humic acids) or marine humic-like substances, hydrophilic DOM (major fraction; ~>50 %),

19 Dissolved Organic Matter in Natural Waters
including organic acids (~14–40 %) such as acetic and formic acid, dicarboxylic
acids (~<6 %, including oxalic, succinic, malonic and maleic acids), pyruvic acid (~<1 %), amino acids (~2 %) including tryptophan-like and tyrosine-like compo-nents, formaldehyde (~2–8 %), acetaldehyde (~5 %), organic peroxides (ROOHs:
see chapter “Photoinduced and Microbial Generation of Hydrogen Peroxide and Organic Peroxides in Natural Waters” for detailed description) (McDowell and Likens 1988; Hellpointner and Gäb 1989; Hewitt and Kok 1991; Guggenberger
and Zech 1993; Sakugawa et al. 1993; Sempéré and Kawamura 1994; Chebbi and Carlier 1996; Williams et al. 1997; Willey et al. 2000, 2006; Ciglasch et al. 2004;
Avery et al. 2006; Kieber et al. 2006; Muller et al. 2008; Miller et al. 2008, 2009;
Santos et al. 2009a, b; Southwell et al. 2010; Zhang et al. 2011; Nichols and Espey
1991; Brassell et al. 1980; Sargent et al. 1981). These studies also show that rainwa-ter mostly consists of low molecular weight organic substances, having MW < 1000 Dalton. Note that factors such as wind speed, storm trajectory and rainwater volume can influence DOM contents in rainwater. The relative importance of these factors depends on the sources of the rainwater constituents (Miller et al. 2008).
The contribution of allochthonous fulvic and humic acids is significantly high
in source waters (streams and rivers), then their contributions decrease during the flow into the downward water ecosystem (lakes, estuaries and oceans) because of three major processes: first, photoinduced and microbial degradation; second, dilu-tion of the source waters with other water bodies; third, high contents of autoch-thonous DOM can decrease the relative contribution of allochthonous fulvic and humic acids in stagnant waters, particularly in lakes, estuaries and oceans.
On the other hand, the contribution of autochthonous DOM including autoch-
thonous fulvic acids of algal or phytoplankton origin, carbohydrates, proteins, amino acids, lipids, organic acids etc. is relatively low in source waters, but sig-nificantly high in lakes and oceans. Autochthonous production of DOM is typi-cally detected in the epilimnion of lake and ocean during the stratification period. A rough estimate shows that the contribution of autochthonous DOM is 0–102 % in lakes and 0–194 % in oceans, which has been discussed in earlier section (Mostofa et al. 2009a; Wigington et al. 1996; Fu et al. 2010; Ogawa and Tanoue
2003; Ogawa and Ogura 1992; Mitra et al. 2000; Yoshioka et al. 2002a; Hayakawa
et al. 2003, 2004; Annual Report 2004; Bade 2004; Sugiyama et al. 2004).
The sterol biomarkers used for identifying DOM sources in water are terrestrial
(b-sitosterol and ergosterol), sewage (5b-coprostanol and epi-coprostanol), phy-toplankton (cholest-5,22-dien-ol, brassicasterol, dinosterol), and marine markers (cholesterol) (McCalley et al. 1981; Mudge and Bebianno 1997; Mudge and Duce 2005; Nichols and Espey 1991). Long-chain C22-C30 alkanols are generally con-sidered to originate from terrestrial plants, while short-chain alkanols have unspec-ified marine, terrestrial and bacterial origins (Brassell et al. 1980; Sargent et al. 1981). From the above contributions to the DOM composition in various sources of waters, it is evidenced that, on average, approximately 80–90 % of bulk DOM in streams, rivers, lakes and oceans is specifically identified as allochthonous ful-vic and humic acids, autochthonous fulvic acids, carbohydrates, proteins, lipids, amino acids, fatty acids, sterols, and organic acids.

20 K. M. G. Mostofa et al.
4.1 Physical and Chemical Properties of DOM
Naturally-originated organic compounds such as humic substances (fulvic and
humic acids) of terrestrial plant origin, autochthonous DOM of algal or phyto-plankton origin, proteins, amino acids, peptides and polysaccharides exhibit, to varying degrees, several major properties (Mostofa et al. 2009a, b; Malcolm
1985; Xue et al. 1995; Mandal et al. 1999; Filella 2008). They are: (i) physically heterogeneous; (ii) polyfunctional, due to the existence of a variety of functional groups and the presence of a broad range of functional reactivity; (iii) polyelec-trolytical, with high electric charge density due to the presence of a large number of dissociated functional groups; (iv) structurally labile, because of their capacity to associate intermolecularly and to change molecular conformation in response to changes in pH, pE, ionic strength, trace metal binding, and so on; (v) polydis-perse in size.
Water Color:
The yellow color in natural waters is due to the occurrence of humic sub-
stances (fulvic and humic acid) and of autochthonous fulvic acids (C-
and M-like) of algal or phytoplankton origin, which absorb light in the blue and ultraviolet (Kalle 1966; Jerlov 1968). These substances were formerly referred to collectively as yellow substances or gelbstoff (Kirk 1976; Kalle 1966). Water color is generally related to the occurrence and contents of these substances in natural waters (Eloranta 1978; Jones and Arvola 1984). Ocean
color is an important feature of water that was recently determined using remote sensing applications (Hopkinson et al. 2002; Morel et al. 2007; Morel and Gentili 2009; Van der Woerd et al. 2011; V olpe et al. 2011; Son et al. 2011). It is mostly due to the effect of autochthonous fulvic acids of algal or phytoplankton origin as well as partly to allochthonous fulvic and humic acids (humic substances). A recent study has shown that autochthonous fulvic acids (C-like and M-like) of lake algal origin under dark incubation can exhibit yel-low color (Mostofa et al. 2009b). Note that autochthonous fulvic acids (C-like and M-like) are characterized based on their similar fluorescence properties to allochthonous fulvic acids (C-like and M-like), which will be discussed in detail in the FDOM chapter (see chapter “Fluorescent Dissolved Organic Matter in Natural Waters”).
Attenuation of Spectral UV Irradiance
DOM is the key factor that controls the downward irradiance flux through the
water column of UV-B (280–320 nm), UV-A (320–400 nm), total UV (280–400 nm) and photosynthetically available radiation (PAR, 400–700 nm) (Kirk 1976; Morris et al. 1995; Siegel and Michaels 1996; Morris and Hargreaves 1997; Tranvik 1998; Bertilsson and Tranvik 2000; Laurion et al. 2000; Markager and
Vincent 2000; Huovinen et al. 2003; Sommaruga and Augustin 2006; Hayakawa

21 Dissolved Organic Matter in Natural Waters
and Sugiyama 2008; Effler et al. 2010). These studies show that UV-B penetration
depths vary from only a few centimeters in highly humic lakes to dozens of meters in the oceans, due to variation in DOM contents. It is also observed that 99 % of the UV-B radiation is attenuated in an approximately 0.5-m water column in the clearest lake for DOC ranging from 408 to 725 μM C and for chlorophyll a rang-ing from 1.6 to 16 μg L
−1 (Huovinen et al. 2003). In the UV-A region at 380 nm,
the corresponding attenuation is limited to the upper one meter.
The absorption coefficients predict that, in a small humic lake (DOC 1100–
1242 μM C), UV-B radiation is attenuated to 1 % of the subsurface irradi-
ance within the top 10 cm water column, whereas UV-A radiation (at 380 nm) penetrates more than twice as deep (maximum 25 cm) (Huovinen et al. 2003). However, in clear lakes with low DOC concentration the contribution of phyto-plankton to UV attenuation can be significant (Sommaruga and Psenner 1997). Any enhancement of photoinduced degradation of DOC by UV radiation and acidification can substantially increase the UV transparency in lakes (Morris and Hargreaves 1997; Vione et al. 2009; Schindler et al. 1996; Yan et al. 1996; Scully
et al. 1997). The consequence is an enhanced penetration of UV radiation into the water column, which can significantly damage aquatic biota. DOM is thus respon-sible for UV attenuation in the water column and for the related protection of aquatic organisms in natural waters.
Aggregation of DOM
Aggregation of fulvic and humic acid (humic substances) can occur at the intra-
molecular (involving a single polymer molecule) or intermolecular (involving multiple chains) levels in aqueous solution (Wershaw 1999; Engebretson and von Wandruszka 1996; Lippold et al. 2008). The interior of the resulting aggre-
gates is relatively hydrophobic, whilst the exterior is more hydrophilic. They can exist in a pseudomicellar form or as micelle-like aggregates in solution, and as membrane-like aggregates on mineral surfaces (Wershaw 1999; Sutton and Sposito 2005; Piccolo et al. 2001). The results of the chemical analysis of
humic acids isolated from natural environments (water, soil, peat, sediments, and sludge from wastewater treatment facilities) demonstrate that the per –
centage elemental composition, the contents of carboxylic groups and of aro-matic phenolic groups is very variable. They range from 33.2 (river) to 60.7 % (Aldrich) of C; 2.25 (river) to 5.4 % (soil) of H; 0.65 (river) to 3.7 % (peat) of N; 34.1 (Aldrich) to 63.8 (river) of O; 0.06 (soil) to 0.10 % (sewage sludge) of
S; 1.0 (river) to 8.1 mmol g
−1 (peat) of carboxylic groups (-COOH), and from
0.36 (bog peat) to 4.4 mmol g−1 of phenolic moieties (ArOH) (Klavins and
Purmalis 2010).
Humic acids behave like surface-active substances when they are added to
solutions, which depend on their origin and molecular properties. Therefore, their surface tension decreases as their concentration increases (Lippold et al. 2008; Klavins and Purmalis 2010; Wershaw 1993; Engebretson and von Wandruszka
1994; Terashima et al. 2004). Humic acids can be significantly modified in their

22 K. M. G. Mostofa et al.
functional groups such as the benzene ring in phenolic structures with the addi-
tion of hydrophilic sulfonic, hydroxyl or trimethylammonium functional groups (Klavins and Purmalis 2010). This effect can be used for the development of biopolymers with surfactant properties (Klavins and Purmalis 2010; Heinze and Liebert 2001). Humic substances might influence plankton food chains in lakes in
two ways (Jones 1992): (i) By altering the physical or chemical environment and thus modifying autotrophic primary production and the dependent food chains; and (ii) By acting as a direct carbon/energy source for food chains.
4.1.1 Redox Behavior of Fulvic and Humic Acids
Fulvic and humic acids (humic substances) can act as reductants and oxidants in
aqueous media (Lovley et al. 1996; Richardson 2007; Wilson and Weber 1979; Nash
et al. 1981; Skogerboe and Wilson 1981; Österberg and Shirshova 1997; Scott et al. 1998; André and Choppin 2000; Steelink 2002; Kuczewski et al. 2003; Shcherbina et al. 2007). They are capable of reducing Fe
3+, Sn4+, V5+ and Cr4+. The +IV oxi-
dation states of the redox-sensitive actinides (e.g. Pa, Np, U, Pu) are stabilized by complexation with fulvic and humic acids. Fulvic and humic acids are thus capable of detoxifying surface water and soils contaminated with toxic organic and inorganic chemicals. Some examples are (i) reduction of metals from toxic valence states to non-toxic states, such as Cr
4+ to Cr3+, V5+ to V4+, or UO 22+ and UO 2OH+ to U4+
(Steelink 2002; Wittbrodt and Palmer 1995; Markich 2002; Freyer et al. 2009); (ii)
reductive cleavage of halogenated hydrocarbons such as trichloroethylene, a common pollutant in soil and groundwater, which can be degraded to ethylene and hydrochlo-ric acid (Steelink 2002); (iii) abiotic reduction of mercury in the presence of a com-peting ion as well as methylation of the carboxylic groups of humic and fulvic acids, which can consume methylmercury (Allard and Arsenie 1991); and (iv) reduction of organic nitro groups to amines. For instance, trinitrotoluene (TNT) is reduced to com-pounds such as aminodinitrotoluene that can form complexes with fulvic and humic acids (Steelink 2002). Note that TNT is an explosive that can migrate to and pollute
groundwater.
On the other hand, it has also been observed that the functional groups in ful-
vic and humic acids can be oxidized, as is the case of catechol moieties (oxidized to quinones), aldehydes (to carboxylic acids), alcohols (to aldehydes or carbox-ylic acids) and so on (Steelink 2002). These redox processes account for the pres-ence of intermediates such as semiquinones in fulvic and humic acids. A typical redox process involving fulvic acids (FA) and humic acids (HA) can be depicted as below (Wilson and Weber 1979; Skogerboe and Wilson 1981):
For instance, SRHA has standard reduction potential E° = 760 ± 6 at pH
5–7. The E° values are variable depending on the pH (Wilson and Weber 1979; (2.1)
FA(ox)+e−+H+=FA(red),E◦=500 mV (FA)and 700 −794 mV (HA)

23 Dissolved Organic Matter in Natural Waters
Skogerboe and Wilson 1981; Matthiesen 1994; Struyk and Sposito 2001). Some
studies also suggest that functional groups such as quinone or quinone-like moie-
ties in fulvic and humic acids are largely responsible for the observed reversible redox behavior in natural waters (Scott et al. 1998; Tratnyek and Macalady 1989; Schwarzenbach et al. 1990; Nurmi and GTratnyek 2002; Cory and McKnight 2005; Macalady and Walton-Day 2009). In addition, fulvic and humic acids can donate electrons photolytically in aqueous media, which can induce the pro-duction of oxidizing agents such as superoxide ion (O
2•−) and hydrogen per –
oxide (H 2O2) (see detailed description in chapter “Photoinduced and Microbial
Generation of Hydrogen Peroxide and Organic Peroxides in Natural Waters”) (Mostofa and Sakugawa 2009; Fujiwara et al. 1993; Baxter and Carey 1983).
The presence of diverse functional groups in the molecular structure of ful-
vic and humic acids is responsible for their redox behavior in waters. The redox behavior of humic acids depends on the redox potential of the aqueous solu-tions as well as on the complexation capacity with multicharged cations in water (Österberg and Shirshova 1997; Struyk and Sposito 2001; Kerndorff and Schnitzer 1980; Zauzig et al. 1993).
4.1.2 Definition and Chemical Nature of Allochthonous Fulvic
and Humic Acids
Allochthonous DOM of vascular plant origin is primarily composed of humic substances (fulvic and humic acids), which are also termed as hydrophobic acids. Stream fulvic and humic acid are therefore vital to understand the nature of the allochthonous DOM, because the chemical composition and optical properties of these substances are greatly altered photolytically and microbially during their transportation after leaching from soil into rivers, lakes or oceans.
Allochthonous Fulvic Acids
Allochthonous fulvic acids can be defined as molecularly heterogeneous and
supramolecular, with molecular weight ranging from less than 100 to over 300,000 Daltons and with the largest fractions ranging less than 50,000. They are opti-cally active, typically refractory to microbial degradation, photolytically reac-tive, biogenic, and yellow-colored. They are also soluble under all pH conditions in water (Ma and Ali 2009; Mostofa et al. 2005b, 2007a; MacFarlane 1978; Dai
et al. 1996; McKnight et al. 1988, 2001; Hayase and Tsubota 1983; Frimmel 2004;
Aiken et al. 1985; Aiken and Malcolm 1987; Aiken and Gillam 1989; Amador
et al. 1989; David and Vance 1991; Allard et al. 1994; Hummel 1997; Fimmen
et al. 2007). Allochthonous fulvic acids in surface waters have relatively low contents of organic N compared to organic C, i.e. a high C:N ratio. This ratio is in the range ~45–202, and standard SRFA (1S101F and 2S101F) have values of 73–78. Allochthonous fulvic acids also have relatively high contents of O and organic P, low contents of S, relatively low aromaticity (17–30 % of total C) and high aliphatic C (63 %) (Malcolm 1985; Wetzel 1983; McKnight et al. 2001;

24 K. M. G. Mostofa et al.
Meyers-Schulte and Hedges 1986; Ma et al. 2001; McIntyre et al. 2005; Frimmel
2004; Aiken and Malcolm 1987; Abbt-Braun and Frimmel 1990; Abbt-Braun et al. 1991; IHSS 2011; Senesi 1990).
Allochthonous fulvic acids are supramolecular structures composed of a
variety of functional groups or components such as benzene-containing car –
boxyl groups, ketones, methoxylate and phenolic groups (catechol-type), carboxylic and di-carboxylic groups, ethers, esters, amides, aliphatic OH, car –
bohydrate OH, –C = C–, hydroxycoumarin-like structures, chromone, xan-
thone, quinones, flavones, O, N, S, and P-atom-containing functional groups attached to aromatic and aliphatic C, indole groups, degraded lignins, and so on (Malcolm 1985; Dai et al. 1996; Frimmel 2004; Allard et al. 1994; McKnight
et al. 1988; Leenheer et al. 1995, 1998, 2001; Brown and Rice 2000; Haiber
et al. 2001; Kujawinski et al. 2002; Lambert and Lankes 2002; Cook et al. 2003; Stenson et al. 2003; Leenheer and Croué 2003; Leenheer 2007; Killops
and Killops 1993). Lignins are complex, high-mass, primarily ether-linked
phenylpropanoid biopolymers including only C, H, and O atoms in their molec-ular structure. They are mostly found in wood cells, whereas the main build-ing blocks for the phenyl portion of lignins are coumaryl, coniferyl, and sinapyl alcohols that vary from plant to plant (Helm 2000; Filley et al. 2002; Lewis and Yamamoto 1990; Christman and Oglesby 1971). The lignin biopolymer
is degraded by fungi and eventually bacteria through different pathways that include depolymerization, demethylation, side-chain oxidation, and aromatic ring cleavage (Lewis and Yamamoto 1990; Nelson et al. 1987; Grushnikov and Antropova 1975; Higuchi 1993; Radnoti de Lipthay et al. 1999; Leonowicz
et al. 2001; Lowe and Bustin 1989).
In humic substances, 60–90 % of the acid groups are carboxylic and the
remainder are phenolic (Leenheer et al. 1995). S-XANES have shown that sulphur is present in humic substances in many different oxidation states: organic sulfides (R–S–R), thiol (–SH), di– and polysulfides (R–S–S–R), sulfoxide (R–SO–R), sul-fone S compounds (R-SO
2-R), sulfonate (HSO 3-R), and sulfate esters (HSO 4-R)
(Frimmel 2004; McKnight et al. 1988; Morra et al. 1997; Xia et al. 1998, 1999;
Schnitzer and Khan 1978).
Depending on the major elemental composition of C, H, O and N disregarding
S, an average empirical formula for fulvic acid has been considered as C 12H12O9N
(Steelink 2002; Leenheer et al. 1998; Paciolla et al. 1998; Schnitzer 1985). Based
on accurate mass measurements, molecular formulas have been assigned to 4626 individual Suwannee River fulvic acids with molecular masses between 316 and 1098 Da, which led to plausible structures consistent with degraded lignin (Leenheer and Croué 2003).
Hummel (Fimmen et al. 2007) has shown that a fulvic molecule (i) contains
on average 5.5 mmoles of carboxyl groups per gram, which corresponds to one carboxylic group per six carbon atoms, or one group per aromatic ring if distrib-uted evenly; (ii) has an average phenolic group content of 1.2 mol per gram, which means one phenolic group per 30 carbon atoms, or only two phenolic groups per

25 Dissolved Organic Matter in Natural Waters
fulvic molecule; and (iii) has hydroxyl and carbonyl groups that, put together,
are as abundant as carboxyl groups (5–7 mmol g−1). Therefore, an average ful-
vic acid molecule (molecular weight 2,000 g mol−1) would have one carboxylic,
hydroxyl or carbonyl group every three carbon atoms. Amino acids, amino sug-ars, ammonium (NH
4+) and nucleic acid bases make up 45–59 % of fulvic acid-N
(Smith and Epstein 1971).
The stable carbon isotope (δ13C = 13C/12C) fractionation of standard SRFA is
–27.6 ‰, while other isolated allochthonous fulvic acids in rivers have [–(25.6–26.4 ‰)] and in lakes have [−(23.02–33.13) ‰]. These data indicate that SRFA are most likely derived from higher plant matter (Thurman 1985a; McIntyre et al.
2005; Senesi 1990; Simpson et al. 2002; Caraco et al. 1998). Note that standard FAs of Elliot Soil I have δ
13C = −25.4 ‰, Elliot Soil II have δ13C = −25.6 ‰,
Pahikee peat I have δ13C = −25.8 ‰. Reference FA of Suwannee River have
δ13C = −27.9 ‰, Pahikee peat I have δ13C = −26.1 ‰, Nordic Lake have
δ13C = −27.8 ‰ (Senesi 1990).
Terrestrial DOM from groundwater, streams, rivers, lakes and sea water
(0 salinity) is confined to a narrow range of δ13C (from –25.3 ‰ to –28.6 ‰),
with 80 % of the values falling within 0.5 ‰ of –27.0 ‰ (Schiff et al. 1997; McCallister et al. 2004; McIntyre et al. 2005; Elder et al. 2000; Nagao et al. 2011; Fry and Sherr 1984). Note that the δ
13C is largely different for fresh
deciduous leaves (–30.4 ‰), it increases in the top soil (–28.9 ‰) and then from –27.8 to –26.4 ‰ in soil. Plant leaves with C3 photosynthesis have δ
13C = –(25.9–29.2 ‰) and soil profiles have δ13C = –(23.8–25.9 ‰). δ13C
has lower values in litter-rich soil DOC [− (26.6–27.7 ‰)] than in litter-lack-
ing soil DOC [approximately − (23–27 ‰)] or terrigenous soil with surface/
forest litter [− (23–27 ‰)], terrestrial leaf OM (− 27 ‰), terrigenous vascular
plant [− (26–30 ‰)], yellow soil profile [− (21.1–24.8 ‰)] or limestone soil
profile [− (23.0–24.1 ‰)] (Tu et al. 2011; McCallister et al. 2004; Elder
et al. 2000; Trumbore et al. 1992; Deegan and Garritt 1997; Stevenson 1982; Richter et al. 1999; Raymond and Bauer 2001b; Cloern et al. 2002; Zhu and Liu 2006; Stenson et al. 2002). Therefore, the origin of allochthonous DOM is
significantly dependent on the types and nature of terrestrial vegetation in soil environments.
The combination of flow path analysis and
14C content of DOC suggests that
the DOC in upland streams is composed of two pools (Schiff et al. 1997). First, the DOC pool is carried to the stream by discharging groundwater. This DOC has been extensively recycled in the soil zone, has low
14C content and proba-
bly has a low proportion of labile functional groups. Although groundwater con-tributions to stream flow are high even during storm events, groundwater DOC concentrations are low. The relative contribution of this older recalcitrant pool is limited by the amount of soluble carbon which elutes through the overlying soil column. The second pool is composed of recently fixed and potentially more microbially labile DOC leached from the A horizon or litter layer. The potential contribution of this second pool is very high especially after leaffall.

26 K. M. G. Mostofa et al.
Allochthonous Humic Acids
Allochthonous humic acids in surface waters can be defined as molecularly het-
erogeneous and supramolecular, with molecular weight ranging from less than 500 to over 300,000 Daltons. The largest fraction is found in the range larger than 300,000 Daltons. They are optically active, typically refractory to microbial degradation, photolytically reactive, biogenic, and yellow-colored organic acids. They are insoluble and form precipitates at pH < 2 (MacFarlane 1978; Hayase and Tsubota 1983; Sutton and Sposito 2005; Steelink 2002; Aiken and Malcolm 1987;
Aiken and Gillam 1989; Schulten and Schnitzer 1998). Allochthonous humic acids of various origin (soil, bog peat, sewerage sludge) have relatively high contents of organic N to organic C, i.e. they have relatively low C:N atomic ratio (8–51). Standard SRHA (1S1011H and 2S101H) have C:N = 44–45. Allochthonous
humic acids also have relatively low contents of O and organic P, high contents of S, relatively high aromaticity (30–40 % of total C) and relatively low contents of aliphatic C (~30–47 %) compared to fulvic acids (Malcolm 1985; Wetzel 1983;
McKnight et al. 2001; Meyers-Schulte and Hedges 1986; Ma et al. 2001; McIntyre
et al. 2005; Frimmel 2004; Aiken and Malcolm 1987; Abbt-Braun and Frimmel 1990; Abbt-Braun et al. 1991; IHSS 2011; Senesi 1990). It has been shown that the contents of aromatic and other functional groups are very variable depending on the different sources of humic acids and their photobiogeochemical changes in natural waters. The aromaticity of humic acids is very low (~15 %) in marine waters (Malcolm 1990).
Allochthonous humic acids have a supramolecular structure composed of a
variety of functional groups (or fluorophores), such as aromatic carboxylic and di-
carboxylic acids, aromatic OH groups including phenols (or catechols) and phe-nolic acids, aliphatic or carbohydrate OH, aldehyde or aliphatic ketones, amide/amino groups, peptides, esters (COOR) or benzene-containing methoxylates, poly-methylenes (–CH
2–), hydroxycoumarin-like structures, chromone, xanthone, qui-
none, O, N, S, and P-atom-containing functional groups attached to aromatic and aliphatic carbon, methylated forms of para-coumaric, ferrulic, vanillic and syringic acids, pyrrole, indole, imidazole and pyridine groups (Malcolm 1985; Sutton and Sposito 2005; Steelink 2002; Lambert and Lankes 2002; Leenheer and Croué 2003;
Stevenson 1982; Schulten and Schnitzer 1998; Laane 1984; Mao et al. 1998; Hu
et al. 2000; Mahieu et al. 2000, 2002; Zang et al. 2000; Kujawinski et al. 2009;
Piccolo 2002; Vairavamurthy and Wang 2002; Abe and Watanabe 2004; Schmidt-
Rohr et al. 2004; Guignard et al. 2005; Fiorentino et al. 2006). A typical humic acid containing 0.2 % reduced sulphur has only 63 μ mol g
−1 of thiol sites (Bloom et al.
2001). Amino acids, amino sugars, ammonium (NH 4+) and nucleic acid bases make
up 46–53 % of the N associated with humic acids (Schnitzer 1985). Depending on the elemental compositions of C, H, O, and N, an empirical formula for humic acids has been proposed as C
10H12O5N and a representative molecular formula as
C72H72O30N4·8H2O (Steelink 2002; Schnitzer and Khan 1978; Paciolla et al. 1998).
The stable carbon isotope (δ13C) fractionation of standard SRHA is –27.7 ‰,
which indicates that they are most likely derived from higher plant matter (IHSS

27 Dissolved Organic Matter in Natural Waters
2011). Note that Standard HAs of Elliot Soil have δ13C = −22.6 ‰; Pahikee peat
have δ13C = −26.0 ‰, and Leonardite have δ13C = −23.8 ‰. Reference HAs
of Suwannee River have δ13C = −28.2 ‰, Pahikee peat have δ13C = −26.3 ‰,
Nordic Lake have δ13C = −27.8 ‰, and Summit Hill soil have δ13C = −26.3 ‰
(IHSS 2011). In addition, carbon isotope composition of dissolved humic and ful-
vic acids shows that the Δ14C values are ranged from −247 to +26 ‰ whilst the
average values are −170 ± 79 ‰ for humic acid and −44 ± 73 ‰ for fulvic
acid (Nagao et al. 2011). This suggests that the residence time of fulvic acid in the
watershed is being shorter than that of humic acid (Nagao et al. 2011).
4.1.3 Definition of Autochthonous Fulvic Acids and Chemical
Nature of Autochthonous DOM
The key autochthonously produced biochemical organic groups or substances (Mostofa et al. 2009a) identified in natural waters can be classified as: autoch-thonous fulvic acids (C-like and M-like) of algal (cyanobacteriam) or phy-toplankton origin; carbohydrates such as uranic acids, amino sugars and neutral sugars including free mono-, oligo-. lipopoly-, exopoly-, homopoly-, and
heteropolysaccharides; nitrogen-containing organic compounds including amino acids, proteins, amines, amides, urea, purines, pyrimidines, peptides, polypep-tides, pyrrole, and indole; lipids, including saturated, monounsaturated, polyun-saturated, branched-chain and odd-chain fatty acids (mostly composed of oleic acid, arachidonic acid, eicosapentanoic acid, linoleic acid, docosahexaenoic acid, cis-vaccenic acid, iso- and anteiso-C
15 and C 17 fatty acids, polyunsaturated C 22
and C 20 fatty acids, high molecular-weight, straight-chain (C 24, C 26, C 28,C30)
fatty acids; organic acids including mono-, di- and tri-carboxylic acids, glycol-late, and hydroxamate; allelopathic compounds. There are also steroidal alcohols (sterols) such as 24-methyl-cholesta-5,24(28)-dien-3ß-ol, 24-ethylcholest-5-en-3ß-ol, cholesta-5,22E-dien-3ß-ol, cholest-5-en-3ß-ol, cholesta-5,22-dien-3ß-ol, 27-Nor-24-methylcholesta-5,22-dien-3ß-ol, 4α,23,24-trimethyl-5α-cholest-22E-en-3ß-ol (dinosterol), 24-methylcholesta-5,22-dien-3ß-ol, 24-ethylcholesta-5,22E-dien-3ß-ol, 24-ethylcholesta-5-en-3ß-ol, 24-ethylcholesta-5,24(28)E-dien-3ß-ol, 24-n-propylcholesta-5,24(28)E-dien-3ß-ol, 3-methyllidene-7,11,15-trimethylhexa-decan-1,2-diol (phytyldiol); vanillyl and syringyl phenols including vanillin, ace-tovanillone, vanillic acid, syringaldehyde, acetosyringone and syringic acid from lignin-derived oxidation products; bisnorhopane and various alkenones such as four polyunsaturated C
37 and C 38 methyl- and ethyl- alkenones, 6,10,14-trimeth-
ylpentadecan-2-one; pigments including melanin, mycosporine-like amino acids (shinorine, palythine, porphyra-334, palythene and usujirene); carotenoids (dia-dinoxanthin, zeaxanthin, myxoxanthophyll, and echinenone); algal toxins (mostly cyanobacterial toxins produced from blue–green algae) including microccystins, nodularins, anatoxins, cylindrospermopsin, and saxitoxins; red tide toxins includ-ing brevetoxins (Parlanti et al. 2000; Mostofa et al. 2009b; Zhang et al. 2009;
Xiao and Wu 2011; Coble 2007; Norrman et al. 1995; Hanamachi et al. 2008;

28 K. M. G. Mostofa et al.
Richardson 2007; Singh and Singa 2002; Miller et al. 2002; Hama et al. 2004;
McCallister et al. 2006; Prince et al. 2008). Most of these autochthonous sub-
stances have been extensively discussed in earlier studies (Mostofa et al. 2009a).
“Autochthonous fulvic acids” of algal or phytoplankton origin are molecularly
heterogeneous, with molecular weight ranging from less than 100 to over 1,898 Daltons. They are optically active, biogenic, highly photoreactive, microbially refractory and yellow-colored organic acids (Mostofa et al. 2009b, Mostofa KMG et al., unpublished data; Zhang et al. 2009; Johannessen et al. 2007; Amon and Benner 1994; McKnight et al. 1991, 1994; Ogawa et al. 2001; Aoki et al. 2004,
2008; Nagai et al. 2005; Williams and Druffel 1987; Fimmen et al. 2007; Barber 1968; Ogura 1972). Autochthonous fulvic acids or DOM in freshwater and seawa-ter have relatively high contents of dissolved organic N compared to organic C, i.e. low C:N atomic ratios (ca. 8–36, but lower in surface waters and higher in deeper waters). They are rich in S, highly aliphatic in nature and have low contents of aromatic carbon (ca. 5–21 % of total carbon) (Wetzel 1983; McKnight et al. 1991, 1994, 1997, 2001; Ogawa et al. 1999, 2001; Meyers-Schulte and Hedges 1986;
Aluwihare et al. 2002; Fimmen et al. 2007; McCallister et al. 2006; Nissenbaum and Kaplan 1972; Carder et al. 1989; Karl et al. 1991; Midorikawa and Tanoue 1996, 1998; McCarthy et al. 1997; Engel and Passow 2001; Carlson et al. 2000;
Church et al. 2002). Autochthonous fulvic acids have higher nitrogen content
(C:N = 8–36) than allochthonous standard fulvic and humic acids (C:N = 44–78).
This may indicate that autochthonous fulvic acids are less refractory than alloch-thonous fulvic and humic acids, probably because autochthonous DOM has fewer aromatic compounds and relatively more proteins and lipids, which decreases its carbon to nitrogen ratio compared to allochthonous DOM (McCallister et al. 2006). Cyanobacteria may contain significant quantities of lipids (fats and oil) which are esters of fatty acids and alcohols that comprise a large group of struc-turally distinct organic compounds including fats, waxes, phospholipids, glycolip-ids etc. (Singh and Singa 2002). The lipids of some cyanobacterial species are
also rich in essential fatty acids such as the C
18 linoleic (18:2ω6) and y-linolenic
(18:3ω3) acids and their C 20 derivatives, eicosapentaenoic acids (20:5ω3) and
arachidonic acid (20:4ω6) (Singh and Singa 2002). These fatty acids are essential components of the diet of humans and animals and are becoming important feed additives in aquaculture (Borowitzka 1988).
Spectroscopic studies of isolated autochthonous fulvic acids show that they
are composed of methylated isomers of hydroxy-benzenes and hydroxy-benzoic acids, aliphatic acids, carbohydrate OH, protein amide and amine groups; they also contain Schiff-base derivatives (–N = C–C = C–N–), fatty acid methyl esters
(heptanedioic acid, octanedioic acid, nonanedioic acid, methyl tetradecanoate, 12-methyl-tetradecanoic acid, 7-hexadecenoic acid, and hexadecanoic acid), N- and S-containing amino and sulfidic functional groups. The latter include 3-(methylthio)-propanoic acid; dimethyl sulfone; N,N-dimethyl-2-butanamine, N-methyl pro-line; N-methyl aniline; 3-piperidinemethanol; 1-methyl-2,5-pyrrolidinedione; 1-methyl-2-piperidinone; caprolactam; 3-ethyl-1,3-dimethyl-2,5-pyrrolidinedione; 2-amino-5,6-dihydro-4,4,6-trimethyl-4 H-1,3-oxazine; 3-ethyl-2,6-piperidinedi-

29 Dissolved Organic Matter in Natural Waters
one; 1,3,5-trimethyl-1,3,5-triazine-2,4,6-trione; 1,3-dimethyl-2,4-pyrimidinedione;
2-methyl-isoindole-1,3-dione; 5-methoxy-2-methyl-indole; 1,3,5-trimethyl-2,4-py-rimidinedione; and 3,3-dimethyl-4-[(2-methoxycarbonyl)ethyl]-2,5-dione-pyrro-lidine (McKnight et al. 1997; Fimmen et al. 2007; Laane 1984; Borowitzka 1988; Wershaw 1992; Xue and Sigg 1993; Xue et al. 1995). The aromatic compounds pre-
sent in autochthonous DOM originate from intracellular quinones in the chloroplasts and mitochondria of algae and bacteria (McKnight et al. 1997; McKnight and Aiken 1998; Klapper et al. 2002).
Algal toxins such as as microcystins and nodularins have high molecular
weight and cyclic peptide structures and are hepatotoxic; anatoxins, cylindrosper –
mopsin and saxitoxins have heterocyclic alkaloid structures. Anatoxins and saxi-toxins are neurotoxic, while cylindrospermopsin is hepatotoxic (Richardson 2007). On the other hand, red tide toxins such as brevetoxins have heterocyclic polyether structures and are neurotoxic. Note that bacteria, algae and their exudates also consist of a mosaic of functional groups such as amino, phosphoryl, sulfhydryl and carboxylic groups. The net charge on the cell wall depends on the pH of the medium (Filella 2008). Algae and bacteria have no lignin-like components in their molecular structure (McKnight et al. 1997; McKnight and Aiken 1998; Opsahl and Benner 1998), thus the low aromaticity of autochthonous fulvic acids can reflect
the lower content of moieties with sp
2-hybridized carbon in cell wall material and
in other components of microbial cells (McKnight et al. 1994).
Algal- or phytoplankton-derived autochthonous fulvic acids can absorb light to
a lesser extent (by approximately 3–5 times) than allochthonous fulvic acids. They show a progressive increase in absorbance with decreasing wavelength that is typi-cal of fulvic acids (McKnight et al. 1991, 1994). However, the autochthonous ful-
vic acids (C-like and M-like) of algae or phytoplankton origin can exhibit higher fluorescence intensity at peak C-region than at peak A-region, which is an opposite behavior compared to allochthonous fulvic acids (C-like and M-like) of terrestrial plant origin (Fig. 1; McKnight et al. 2001; Mostofa et al. 2009b). Autochthonous fulvic acids can persist with ages up to 3,000 yr in the desert lakes in Antarctica (McKnight et al. 1991, 1994).
The stable carbon isotope (δ
13C) fractionation of autochthonous DOM of algal
or phytoplankton origin ranges from −17.2 to 23.7 ‰ in lake and marine envi-ronments (Thurman 1985a; Raymond and Bauer 2001a; Nissenbaum and Kaplan 1972). The δ
13C values of algae or phytoplankton shows high variation in fresh-
water [−(18.3–34.6 ‰)] and sea water [−(18–24.2 ‰)] (Mostofa KMG et al.,
unpublished data; McCallister et al. 2004; McKnight et al. 1997; Fry and Sherr 1984; Anderson and Arthur 1983; Sigleo and Macko 1985; Yoshioka et al. 1989;
Currin et al. 1995; Yoshioka 1997; Lehmann et al. 2004). In addition, δ
13C shows
high variations between benthic microalgae [−(12–18 ‰)]; benthic marsh micro-algae [−(23.7–27.7 ‰)]; C-4 salt marsh plants [−(12–14 ‰)]; C-3 freshwater/brackish marsh plants [−(23–26 ‰)]; submerged macrophytes [−(21.7–22.2 ‰)]; emergent macrophytes (−26 ‰); marsh macrophytes [−(23.3–28.9 ‰)]; marsh OM [−(22.3–26.4 ‰)]; and freshwater grass leachate such as Peltandra virgi-nica [−(29.6 ‰)] (McCallister et al. 2004; Raymond and Bauer 2001a, c; Caraco

30 K. M. G. Mostofa et al.
et al. 1998; Fry and Sherr 1984; Currin et al. 1995; Sullivan and Moncreiff 1990).
Depending on the origin of DOM from these algae and plants, there can be found variable carbon isotope ratios for DOM in natural waters.
The autochthonous DOM of algal or phytoplankton origin is usually very suit-
able for bacterial use, as suggested by the pattern of increased bacterial produc-tion with increased primary production (Cole et al. 1988). Autochthonous DOM is in fact relatively labile (Søndergaard and Middelboe 1995; Kirchman et al. 1991). However, autochthonously derived DOC may become persistent over time (Ogawa et al. 2001; Fry et al. 1996; Tranvik and Kokalj 1998). Laboratory studies have shown that natural assemblages of marine bacteria become rapidly able (in <48 h) to utilize labile compounds (glucose, glutamate) and produce refractory DOM that can persist for more than a year (Ogawa et al. 2001). It has also been shown that only 10–15 % of the bacterially derived DOM is identified as hydrolysable amino acids and sugars, which is a characteristic nature of marine DOM (Ogawa et al. 2001). Moreover, the higher concentrations of DON observed in total DOM during the summer period than in winter (Fellman et al. 2009; Vazquez et al. 2011) are most likely accounted for by the produced autochthonous DOM in natural waters.
4.2 Molecular Size Distribution of DOM
The molecular size distribution of DOM is significantly variable in natural waters (Table 1). One of the techniques for isolating DOM in natural waters is tangen-
tial flow ultrafiltration (also called cross-flow ultrafiltration). The results show that the contributions of the various fractions to total DOC are 21–65 % for the frac-tion <1 kDa, 44–68 % for <5 kDa, 57–65 % for <10–12 kDa. Moreover, they are 41 % for 1–30 kDa, 32–56 % for 1 kDa–0.1 μ m, 67–84 % for 1 kDa–0.45 μ m,
and 0.1–16 % for 0.1–0.45 or 0.1-GF/F μ m in rivers (Table 1 ) (Yoshioka et al.
2007; Guéguen et al. 2002, 2006; Martin et al. 1995; Mannino and Harvey 2000;
de Zarruk et al. 2007; Wu and Tanoue 2001; Wu et al. 2003; Waiser and Robarts
2000; Huguet et al. 2010; Carlson et al. 1985). In lakes, the relative abundances of various DOM fractions are 42–73 % for <1 kDa, 54–79 % for <5 kDa, 21–43 % for 5 kDa–0.1 μ m, and 0–2 % for 0.1–0.45 μ m (Table 1 ) (Yoshioka et al. 2007;
Guéguen et al. 2002; Wu and Tanoue 2001; Wu et al. 2003; Waiser and Robarts
2000). In estuaries or lagoons, the contributions are 26–98 % for <1 kDa, 11–25 % for 1–3 kDa, 63–75 % for <10 kDa, 25–31 % for 1–30 kDa, 2–45 % for 1 kDa–0.2 μm, 22–48 % for 3 kDa–0.2 μ m, 14–20 % for 30 kDa–0.2 μ m, and 1–2 %
for 30 kDa–0.2 μ m (Table 1 ) (Hagedorn et al. 2004; Mannino and Harvey 2000;
Guéguen et al. 2002; Waiser and Robarts 2000; Huguet et al. 2010). In coastal and open oceans, the contributions of the relative DOM fractions are 30–85 % for <1 kDa (30–70 % in coastal waters, 49–85 % in the open ocean), 23–53 % for the fraction between 1 kDa and 10 kDa, 3–19 % for the fraction between 10 kDa and 0.1–0.2 μ m, 15–70 % for the fraction between 1 kDa and 0.2 μ m, 85 % for
the fraction between 1.8 kDa and 0.2 μ m (Table 1 ) (Buesseler et al. 1996; Druon

31 Dissolved Organic Matter in Natural Waters
Table 1 Molecular size distribution of the fractionated DOM in natural waters
Samples Contribution % of molecular size distribution of DOM References
<1 kDa <3–3.5 kDa <5 kDa <10–12 kDa <30 kDa <0.1 μm <0.2–0.45 μm
or 0.1-GF/F
Soils
Soil DOM, collected
French agriculture– ~50 – 23 – – – de Zarruk et al. (2007)
RiversRivers, Lake Biwa watershed – – 58–68 – – 32–40 0.2–16.0 Yoshioka et al. (2007)
Rivers, Lake Baikal watershed – – 44–56 – – 43–56 0.1–0.5 Yoshioka et al. (2007)
Delaware river 51 – – – 41 – 2 Mannino and
Harvey (2000)
Pearl River, Guangzhou
section, China65 – – – – – –
Vistula river, Poland 21–64 – – – – – 36–86 Guéguen et al. (2002)
Channel fresh water, Venice
Lagoons, Itali (n = 3)– – – 57–65 – – 13–26 Martin et al. (1995)*
Yukon river, Canada – – – – – – 67–84 (>1 kDa) Guéguen et al. (2006)
LakesLake Biwa (2.5 m, n = 2) – – 68–77 – – 22–28 0.8–1.1 Yoshioka et al. (2007)
Lake Biwa (70 m, n = 2) – – 68–76 – – 23–32 0.2–0.4 Yoshioka et al. (2007)
Lake Biwa (2.5 and 20 m
depth: n = 3)– – 54–59 – – 40–43 2 Wu and Tanoue (2001)
Lake Biwa (2.5 depth: n = 2) – – 55–58 – – 40–43 2 Wu et al. (2003)
Lake Biwa (70 m depth: n = 1) – – 69 – – 30 1 Wu and Tanoue (2001)
Lake Biwa (70 m depth: n = 1) – – 69 – – 30 1 Wu et al. (2003)
Lake Baikal (2 m, n = 2) – – 70–77 – – 23–29 0–0.2 Yoshioka et al. (2007)
Lake Baikal (200 and
1400 m, n = 4)– – 72–79 – – 21–27 0.1–0.5 Yoshioka et al. (2007)
(continued)

32 K. M. G. Mostofa et al.Samples Contribution % of molecular size distribution of DOM References
<1 kDa <3–3.5 kDa <5 kDa <10–12 kDa <30 kDa <0.1 μm <0.2–0.45 μm
or 0.1-GF/F
Redberry lake 73 – – – – – – Waiser and
Robarts (2000)
Creeks (Oscar and Trout pond) 55–61 – – – – – – Waiser and
Robarts (2000)
Lake Geneva, France 42–64 – – – – – 36–58 Guéguen et al. (2002)
Estuaries or Lagoons
Venice Lagoons, Itali (n = 5) – – – 63–75 – – 14–20 Martin et al. (1995)
Delaware Estuary 71–74 – – – 25–31 – 1–2 Mannino and
Harvey (2000)
Gironde Estuary: surface water,
French Atlantic coast41–47 13–17 – – – – 22–32 Huguet et al. (2010)
Gironde Estuary: deep water,
French Atlantic coast41–47 11–19 – – – – 31–43 Huguet et al. (2010)
Seine Estuary: surface water,
French Atlantic coast26–56 12–24 – – – – 22–48 Huguet et al. (2010)
Seine Estuary: deep water,
French Atlantic coast35–48 20–25 – – – – 22–34 Huguet et al. (2010)
Adour Estuary, France 55–98 – – – – – 2–45 Guéguen et al. (2002)
OceansGalveston Bay 30–37 – – – – – 63–70 Santschi et al. (1995)
Chesapeake Bay and
Galveston Bay39–41 – – 46–53 – – 7–11 Guo and
Santschi (1997a)
Middle Atlantic Bight 65–70 – – 23–30 3–11 Guo et al. (1996)
Middle Atlantic Bight 51–59 – – – – – 41–49 Santschi et al. (1995)
Gulf of Mexico and
Middle Atlantic Bight55–65 ~24 – 7–14 – – 4–7 Guo et al. (1995)Table 1 (continued)
(continued)

33 Dissolved Organic Matter in Natural Waters
Samples Contribution % of molecular size distribution of DOM References
<1 kDa <3–3.5 kDa <5 kDa <10–12 kDa <30 kDa <0.1 μm <0.2–0.45 μm
or 0.1-GF/F
Gulf of Mexico 47–66 – – – – – 34–53 Santschi et al. (1995)
Gulf of Mexico 55 – – 35 – – 10 Guo et al. (1994)
Open Ocean deep water
(off Hawaii: ~600 m deep)85 – – – – – 15 Mopper et al. (1996)
North Pacific Ocean,
subarctic region (8–59 m)49–62 – – 26–33 10–19 – Midorikawa and
Tanoue (1998)
Open North Pacific Ocean
(22°45′N, 158°00′W)67–78 – – – – – 22–33 Benner et al. (1992)
North Atlantic surface waters 50–70 – – – – – 34 Carlson et al. (1985)
Northwestern Pacific Ocean
surface and deep waters– – – – – – 85 (>1.8 kDa) Sugimura and
Suzuki (1988)
North Pacific Ocean:
deep waters65 – – – – – – Guo and
Santschi (1996)
North Pacific Ocean:
deep waters57 – – – – – – Buesseler et al. (1996)Table 1 (continued)

34 K. M. G. Mostofa et al.
et al. 2010; Midorikawa and Tanoue 1998; Carlson et al. 1985; Sugimura and
Suzuki 1988; Guo et al. 1994, 1995, 1996; Santschi et al. 1995; Guo and Santschi
1996, 1997a; Mopper et al. 1996). These results demonstrate that the contribution
of the lower MW fraction (<1–10 kDa) is relatively low in rivers and that it sig-nificantly increases in lakes, coastal waters and the open ocean. Comparison of molecular fractions between surface (epilimnion) and deep (hypolimnion) waters shows that the molecular size fraction of <1–5 kDa in deep water is often more important than in the surface waters of lakes and oceans (Table 1 ) (Yoshioka et al.
2007; Wu and Tanoue 2001; Wu et al. 2003; Mopper et al. 1996). It is suggested
that either microbial degradation of DOM or new releases of DOM from micro-bial respiration of organic matter in deeper waters are responsible for the high contents of the low molecular size fractions of DOM in natural waters. An addi-tional implication is that significant microbial or biological degradation of DOM
and organic matter occurs in deep waters. The high percentage of colloidal DOC or colloidal organic carbon included in the >1 kDa to 0.45 μ m range suggests that
colloids are the predominant phase in bulk DOC transported by rivers (Guéguen et al. 2006; Benner and Hedges 1993; Guo and Santschi 1997b; Guéguen and Dominik 2003).
The optical and chemical characteristics of the molecular size fractions of
DOM show that truly dissolved DOM (<1–10 kDa) includes fulvic acid (59–96 % on the basis of fluorescence), total hydrolyzed amino acids (51–63 %), tryp-tophan (free tryptophan has a molecular weight of 0.2 kDa) and total dissolved carbohydrates (10–20 %). In contrast, the DOM fraction between >1–10 kDa and 0.2–0.45 μm or 0.1-GF/F includes fulvic acid (5–22 % on the basis of fluores-
cence), total dissolved carbohydrates (80–90 %) and total hydrolyzed amino acids (29–42 %). The DOM fraction of 0.1 μm-GF/F (0.45–0.7 μm) includes protein-like or tryptophan-like or bacterial cells or phytoplankton cells, total hydrolyzed amino acids (7–11 %) and fulvic acid (2–8 % on the basis of fluorescence) (Liu et al. 2007; Guéguen et al. 2006; McCarthy et al. 1996; Midorikawa and Tanoue 1998; Wu and Tanoue 2001; Wu et al. 2003; Pakulski and Benner 1992; Skoog and
Benner 1997; Boehme and Wells 2006). The contributions to the molecular size
fractions of sedimentary fulvic acid extracted from Tokyo Bay sediment samples are 44.8 % for <1 kDa, 3.5 % for 10 kDa, 31.8 % for 50 kDa, 14.6 % for 100 kDa
and 5.3 % for 300 kDa. The corresponding contributions of humic acid are 2.4 % for <1 kDa, 0.8 % for 10 kDa, 5.3 % for 50 kDa, 16.1 % for 100 kDa and 75.4 % for 300 kDa (Hayase and Tsubota 1983, 1985). This suggests that allochthonous
fulvic acid is mostly composed of low molecular size fractions (<1–10 kDa) whilst allochthonous humic acid is mostly composed of high molecular size fractions, >300 kDa (Hayase and Tsubota 1983, 1985; Rashid and King 1969; MacFarlane
1978). Therefore, molecular size fractions could be a useful indicator to distin-guish between fulvic and humic acids in DOM in a variety of natural waters.
These results also imply that allochthonous fulvic acid of terrestrial origin or
the autochthonous fulvic acid (C-like) of algal or phytoplankton origin can primar –
ily undergo photoinduced and microbial in situ degradation, which can decrease the molecular size and increase as a consequence the low molecular size fraction

35 Dissolved Organic Matter in Natural Waters
of DOM (Yoshioka et al. 2007; Amon and Benner 1994; Corin et al. 1996; Amador
et al. 1989; Leenheer and Croué 2003; Opsahl and Benner 1998; Boehme and Wells 2006; Mopper et al. 1991; Senesi et al. 1991; Allard et al. 1994; Benner and
Biddanda 1998; Mopper and Kieber 2002). The autochthonous fulvic acid (C-like) of
algal or phytoplankton origin can show the fluorescence excitation-emission (Ex/Em) maxima of peak C in a longer wavelength region (Ex/Em = 340–370/434–480 nm),
whilst the autochthonous fulvic acid (M-like) can show its Ex/Em maxima in a shorter wavelength region (290–330/358–434 nm) compared to allochthonous fulvic acids (standard SRFA at Ex/Em = 325–345/442–462 nm in Milli-Q and Seawater)
(Parlanti et al. 2000; Mostofa et al. 2009b; Zhang et al. 2009; Vähätalo and Järvinen
2007; Yamashita and Jaffé 2008; Nakajima 2006; Murphy et al. 2008; Balcarczyk
et al. 2009). Note that autochthonous fulvic acids (C-like and M-like) are defined
on the basis of the similarity with the fluorescence properties of allochthonous ful-
vic acids (C-like and M-like) for both freshwater and marine environments (for a
detailed explanation see the FDOM chapter: “Fluorescent Dissolved Organic Matter in Natural Waters”).
Humic-like fluorescence is a key component in DOM size fractions between
~15 and 150 kDa. A bathychromic shift (blue shift) of the humic fluorescence
peak is often detected with decreasing molecular size, and interestingly the maxi-
mum in humic fluorescence moves to lower excitation and emission wavelengths in estuarine waters (Boehme and Wells 2006). Blue-shift phenomena are generally observed in field studies (Coble 1996; Mostofa et al. 2005a, b, 2007a, b; Moran
et al. 2000; Burdige et al. 2004; de Souza-Sierra et al. 1994; Komada et al. 2002).
The molecular size distribution of DOM plays significant roles in various kinds
of physical, photoinduced and biological processes in natural waters. They are
listed below.
(i) The bioreactivity of POM and DOM decreases along a continuum of larger to
smaller sizes. Diagenetic processes lead to the formation of structurally com-plex LMW compounds that are more resistant to biodegradation (Amon and
Benner 1994, 1996; Hama et al. 2004; Mannino and Harvey 2000; Harvey
and Mannino 2001; Benner 2002; Loh et al. 2004; Zou et al. 2004; Seitzinger
et al. 2005; Kaiser and Benner 2009). This hypothesis is termed as size-
reactivity continuum model and is based on the results of size-fractionation experiments that demonstrate that bacterial utilization of (HMW) DOM is typically higher compared to (LMW) DOM (Amon and Benner 1994). It has
also been shown that neutral sugars and amino sugars are considerably more
bioreactive than amino acids in all organic matter size fractions of DOM in deep mesopelagic waters (Kaiser and Benner 2009). Furthermore, nonspecific enzyme reactions can lead to secondary products that are resistant to degrada-
tion (Ogawa et al. 2001). Products of such enzymatic degradations may not
resemble the structure of the original compounds, thereby reducing enzymatic
recognition and further biodegradation. In addition, size can affect the biore-
activity of individual organic matter fractions. Colloidal organic matter, which is part of HMW DOM, is much less accessible to bacteria than particles larger

36 K. M. G. Mostofa et al.
than a few μm because it occupies a minimum between two different trans-
port regimes (Kaiser and Benner 2009; Kepkay 1994; Wells and Goldberg
1993). In fact, Brownian motion dominates transport of smaller colloids to bacteria, whilst larger particles are primarily transported to bacteria by turbu-lent shear (Kepkay 1994).
(ii) Photoinduced and microbial processes that involve DOM, including fulvic and humic acids, can produce biologically labile LMW organic substances (e.g. organic acids) in natural waters (Moran and Zepp 1997; Carrick et al. 1991; Kieber et al. 1989, 1990; Corin et al. 1996; Mopper et al. 1991; Allard
et al. 1994; Mopper and Stahovec 1986; Backlund 1992). These LMW organic compounds are important intermediates of the conversion of organic substances such as carbohydrates, fats and proteins into CH
4 and CO 2 in
aqueous media (Smith and Oremland 1983; Evans 1998; Xiao et al. 2009; Wellsbury and Parkes 1995).
(iii) The absorption of natural sunlight is greatly dependent on the molecular size of DOM and has a high biogeochemical importance in natural waters. For example, fulvic and humic acids (humic substances) can absorb both vis-ible and UV radiation (Kieber et al. 1990; Kramer et al. 1996; Sadtler 1968; Strome and Miller 1978). Many low molecular weight organic acids photo-
generated from large CDOM or FDOM can only absorb in the UV-C range, with no absorption of UV-B, UV-A or visible radiation (Carrick et al. 1991;
Kieber et al. 1990; Mopper et al. 1991; Sadtler 1968). Further details are
provided in the DOM degradation chapter (see chapter “Photoinduced and
Microbial Degradation of Dissolved Organic Matter in Natural Waters”).
4.3 Autochthonous Fulvic Acids and their Differences
with Allochthonous Fulvic Acids
The key component of autochthonous DOM is variously termed as marine humic-
like substances (Coble 1996), sedimentary fulvic acid (Hayase et al. 1987, 1988) or
marine fulvic acids (Malcolm 1990), which is contradictory. However, recent stud-ies show that the two fluorescent components are primarily produced under either photoinduced or microbial respiration (or assimilation) of algal (phytoplankton) biomass (Mostofa et al. 2009b; Zhang et al. 2009; Stedmon and Markager 2005a). PARAFAC modeling of EEM spectra of algal-originated DOM suggests that the fluorescence peaks and the images of the first fluorescent component are similar to those of allochthonous fulvic acid (Fig. 1 ). On the other hand, the fluorescence
peaks and the images of the second fluorescent component are similar to those of marine humic-like substances (Coble 1996). However, the fluorescence intensity and the peak positions of the first fluorescent component are quite different from EEM spectra of standard fulvic acid, which justifies their being denoted with a new name. To avoid the difficulties of indicating the two algal-originated fluorescent compo-nents and considering the similarities of their EEM images with allochthonous

37 Dissolved Organic Matter in Natural Waters
fulvic acid, it is suggested to denote the first and the second fluorescent component
as ‘autochthonous fulvic acid (key component)’ and ‘autochthonous fulvic acid (minor component)’, respectively. These names could be useful to denote the two fluorescent components originated from algae or phytoplankton in fresh- and marine waters in future research studies. The differences between autochthonous fulvic acids and allocthonous fulvic acids and their identification are extensively discussed in the next chapter, ‘Fluorescent dissolved Organic Matter in Natural Waters’.
5 Measurement, Distribution and Sources of DOM
in Natural Waters
Measurement of DOM
DOM is generally determined as dissolved organic carbon (DOC) concentration,
because of the predominant presence of organic carbon in all dissolved organic substances included in bulk DOM. The amount of DOC in natural waters is deter –
mined using a high-temperature catalytic oxidation (HTCO) method developed by Sugimura and Suzuki (Sugimura and Suzuki 1988). This technique is very precise and rapid for the determination of non-volatile DOM in concentrations between 0 and 2000 μ M, compared to conventional wet chemical oxidation meth-
ods (Menzel and Vaccaro 1964; Jonathan 1973). In the HTCO method (Sugimura and Suzuki 1988), the oxidation of DOM in water is carried out on a platinum catalyst at 680 °C under an oxygen atmosphere after the sample has been freed of inorganic carbon. The concentration of CO
2 generated is measured with a non-
dispersive IR gas analyzer. The determination can be carried out with a precision of ±2 % using a sample volume of 100–200 μ l.
Methodology for HTCO (Sugimura and Suzuki 1988): After collection of
water samples using polycarbonate bottles, water is filtered with precombusted (450 °C) glass-fiber or any other filters (0.1–0.7 μm size). Triplicate samples (15 ml) are stored in brown glass bottles (30 ml in volume). 25 μl of 6 N HCl solution is added to remove dissolved inorganic carbon (DIC). These bottles are sealed with Teflon-coated butyl-rubber stoppers and aluminum caps and stored in a freezer (−40 °C). There is need to analyze the samples as soon as possible. For sample measurement, DIC is firstly removed by bubbling the brown bottles with pure air for approximately 15 min. After removing DIC, 200 μl of the water sam-ple is injected into a TOC analyzer (e.g. TOC-5000A, Shimadzu, Kyoto, Japan). Note that analytical blanks for the DOC measurement originating from the instru-ment (system blank) and from pure water (e.g. Milli Q TOC, Millipore) are on average in the range of 2–4 μM C and 6 μM C, respectively (Yoshioka et al. 2002a). The system blank is determined during sample measurement according to the instrument software of the TOC 5000A. The system blank is generally used for the correction of DOC concentration for samples. Potassium hydrogen phthalate is generally used as a standard for calibration.

38 K. M. G. Mostofa et al.
5.1 Distribution and Sources of DOM in Natural Waters
DOM Contents in Stream, Rivers, Groundwater and Rainwater
DOC concentrations are very variable in different upstream locations of the
world (Table 2 ). Relatively low values such as 7–970 μ M C are found in
Asia (Mostofa et al. 2005a, b, 2010, Mostofa KMG et al., unpublished data;
Hayakawa et al. 2004; Konohira and Yoshioka 2005) and 17–3300 μ M C in
North America (Table 2) (McKnight et al. 1993, 2001; V olk et al. 1997; David
and Vance 1991; Fellman et al. 2009; Eckhardt and Moore 1990; Dosskey
and Bertsch 1994; Wahl et al. 1997; Cory et al. 2004; Meier et al. 2004;
Fahey et al. 2005; Raymond and Saiers 2010). Value found in Europe are a bit higher (21–6250 μ M C) (Stedmon et al. 2007b; Worrall et al. 2004a; Evans
et al. 2006; Chapman et al. 2001; Monteith and Evans 2005; Gielen et al. 2011). Stream DOM is mostly released from the leaching of ground water in high mountain areas that in Asia are densely shaded by coniferous-mixed forests or typical grassland. In Europe-North America, the major sources of stream DOM are riparian vegetation, woodland streams (major sources of detritus), wetlands, swamps, and peat-land.
DOC concentrations in rivers vary in different locations of the world (Table 2).
It has been found 32–2429 μM C in Asia (Mostofa et al. 2005b, 2007a, 2010,
Mostofa KMG et al., unpublished data; Yoshioka et al. 2002b, 2007; Ittekkot et al.
1985; Safiullah et al. 1987; Cauwet and Mackenzie 1993; Tao 1996; Zhang 1996;
Kao and Liu 1997; Zhang et al. 1999; Gao et al. 2002; Nagao et al. 2003; Ishikawa et al. 2006; Yue et al. 2006; He et al. 2010); 83–833 μM C in Africa (Martins and Probst 1991); ~50–3917 μM C in Europe (Vazquez et al. 2011; Eisma et al. 1982; Cadée 1987; Meybeck et al. 1988; Rostan and Cellot 1995; Elbe 1997; Lara et al. 1998; Veyssy 1998; Duff et al. 1999; Abril et al. 2000, 2002; Baker 2001,
2002; Brodnjak-V ončina et al. 2002; Guéguen et al. 2002; Baker and Spencer 2004; Kaiser et al. 2004; Romani et al. 2004); 40–4167 μM C in North America (Wu et al. 2005; Xie et al. 2004; McKnight et al. 2001; Alberts and Takács 1999;
Guéguen et al. 2006; Morel and Gentili 2009; Raymond and Bauer 2001b; Haines 1979; Newbern et al. 1981; Spiker 1981; Alberts et al. 1984; Findlay
et al. 1991; Perry and Perry 1991; Prahl et al. 1998; Crandall et al. 1999; Davis et al. 2001; Biddanda and Cotner 2002; Repeta et al. 2002; Wang et al. 2004; Zanardi-Lamardo et al. 2004; Schwede-Thomas et al. 2005; See and Bronk 2005; Stepanauskas et al. 2005; Osburn et al. 2009); and ~108–7500 μM C in Latin America (Raymond and Bauer 2001b; Richey et al. 1990; Depetris and Kempe
1993; Daniel et al. 2002). These results generally show that DOC concentra-tions are relatively low in Asian and African Rivers and relatively high in Europe, North and South America Rivers. The major sources of DOC in Asian Rivers are natural ones such as leaching of groundwater in mountainous areas covered by coniferous-mixed forests, deciduous conifer forest, grassland, irrigated grass-land, and swamps, but also anthropogenic sources such as urban sewerage, indus-trial and agricultural activities. In African Rivers, the major sources of DOC are mostly from the typical rain forest belt, leaching and heterotrophic processes of

39 Dissolved Organic Matter in Natural Waters
Table 2 Distribution of dissolved organic carbon (DOC) concentrations in a variety of natural waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Streams: Asia
Upstreams, Lake Biwa watershed (n = 8) Japan (3174) 7–110 – Mostofa et al. (2005a)
Upstreams, Lake Biwa watershed, Okutama and
Uryu regions (n = 35)Japan – 12–280 – Konohira and Yoshioka
(2005)
Upstream, Kurose River, Hiroshima Prefecture
(n = 2)Japan – 47–239 – Mostofa et al. (2005b)
Upstream, Ohta River, Hiroshima Prefecture
(n = 1)Japan – 59–67 – Mostofa et al. (2005b)
Upstream, Lake Xingyun basin (n = 1) 24șN – ~50 – Hayakawa et al. (2004)
Upstream sites, Nanming River 26șN – 50–100 – Mostofa et al. (2010)
Branches of Upper Region, Yellow River (Huang
He) (n = 13)China – 174–970 – Mostofa KMG et al.,
(unpublished data)
Europe
Upstreams (Warkworth and Afon Hafren) (n = 2) UK – 21–1000 – Worrall et al. (2004a)
Forest streams, Denmark 55șN – 274–1051 – Stedmon and Markager
(2005b)
Fuirosos, forested stream Spain – 150 – Vazquez et al. (2011)
De Inslag forest, Brasschaat, Belgian Campine
regionBelgian (51șN) – ~1000–6250 – Gielen et al. (2011)
Upstreams, The United Kingdom Acid Waters
Monitoring Network (n = 11)UK (2.1–14.1) 100–1075 – Monteith and Evans (2005),
Evans et al. (2006)
Streams, Scotland Scotland – 400–933 – Chapman et al. (2001)
Stream, boreal watershed, in northern Sweden Sweden (2940) 1308–2192 – Bertilsson et al. (1999)
North AmericaUpstream rivers (Hubbard Brook) USA – 158–717 – Fahey et al. (2005)
Upstream Rivers (National Park) USA – 17–367 – Cory et al. (2004)
(continued)

40 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Streams, central Maine (n = 11) USA – 125–1141 – David and Vance (1991)
Stream, Fourmile Branch watershed, South
CarolinaUSA (12.6) 442 – Dosskey and Bertsch (1994)
Streams, Colorado Rocky Mountains USA – 108–433 – McKnight et al. (1993, 2001)
Streams, Dog Creek and Oyster Creek, forested
watershed (n = 2)USA – 933 ± 167–
2208 ± 333– Wahl et al. (1997)
Tongass National Forest, stream USA – 275–358 – Fellman et al. (2009)
Upstream, White Clay Creek USA – 67–867 – V olk et al. (1997)
Upstreams, forest land watershed (n = 30) USA (1.9–226) 75–358 – Raymond and Saiers (2010)
Stream water, New Jersey Pine Barrens USA – 2100 – Meier et al. (2004)
Upstreams, south-central Ontario (n = 11) Canada – 225–900 – Wu et al. (2005)
Streams (n = 42) Canada – 290–3300 – Eckhardt and Moore (1990)
Groundwaters
Groundwater (Tubewell waters) Bangladesh – 17–424 – Anawar et al. (2002)
Groundwater, BaiCheng City and main west
Liaohe river, North-east China (n = 2)China – 49–371 – Mostofa KMG et al.,
(unpublished data)
Groundwater, Lake Biwa basin, Japan Japan – 16–328 (40–
100 m)– Mostofa et al. (2007a)
Groundwater, Tomago sand beds, Newcastle Australia – 183–842 – McIntyre et al. (2005)
Groundwater, Germany Germany – 42–1400 (1–24 m) – Buckau et al. (2000)
Groundwater, Germany Germany – 6117–15333
(35–137 m)– Buckau et al. (2000)
Soil water, De Inslag forest, Brasschaat,
Belgian Campine regionBelgian (51șN) – ~1000–6250 – Gielen et al. (2011)
Groundwater, boreal watershed, in northern
SwedenSweden – 200–850 – Bertilsson et al. (1999)Table 2 (continued)
(continued)

41 Dissolved Organic Matter in Natural Waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Groundwater (Shallow and deep wells; groundwa-
ter spring), Norton-in-HalesUK (38) 467 ± 375–
2708 ± 1558– Bradley et al. (2007)
Groundwater, Suwannee River basin USA – 8–2333 – Crandall et al. (1999),
Schwede-Thomas et al.
(2005)
Shallow groundwater (1.6 m), New Jersey Pine
BarrensUSA – 1558–2583 – Meier et al. (2004)
Groundwater, various types of aquifers USA – 42–8333 – Thurman (1985a)
Groundwater, Cape Cod, USA USA – <167 Pabich et al. (2001)
Groundwater, Amazon Basin Brazil – 100–3000 – Richey et al. (2002)
Groundwater, Okavango Delta Botswana – 1108 ± 217–
14167 ± 6333– Mladenov et al. (2007)
Soil water (upper soil horizons) North America;
Europe– 1667–7500 – Michalzik et al. (2001)
Soil water (lower soil horizons) North America;
Europe– 167–2917 – Michalzik et al. (2001)
Rainwater
Rainwater, upstream regions of Yellow River
(n = 2)China – 93–1784 – Mostofa KMG et al.,
(unpublished data)
Rainwater (n = 483), Northern China China – 25–3675 – Pan et al. (2010)
Rainwater (n = 13), Guangzhou city China – 78–694 – Xu et al. (2008)
Rainwater, Brazilian savanna Brazil – 217 – Ciglasch et al. (2004)
Rainwater (summer), Westwood, Los Angeles USA – <1908 – Sakugawa et al. (1993)
Rainwater (winter), Westwood, Los Angeles USA – 17–758 – Sakugawa et al. (1993)
Rainwater (n = 120), University of North Carolina,
Wilmington campusUSA – 4–379 – Kieber et al. (2006)
(continued)Table 2 (continued)

42 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Rainwater (n = 4), University of North Carolina,
Wilmington campusUSA – 32–105 – Kieber et al. (2007)
Rainwater, mixed forests in New Hampshire and
New YorkUSA – 92–158 – Likens et al. (1983)
Rainwater (n = 18), University of North Carolina,
WilmingtonUSA – 77 ± 20 – Likens et al. (1983)
Rainwater (n = 13), Wilmington, North Carolina,
USAUSA – 12–461 – Southwell et al. (2010)
Rainwater, UNCW campus, Southeastern, North
CarolinaUSA – 81 – Avery et al. (2006)
Rainwater, Wilmington, NC USA – 5–238 – Miller et al. (2008)
Rainwater, Fichtelgebirge Germany – 333–633 – Guggenberger and Zech
(1993)
Rainwater, Meteorological Station, University of
AveiroPortugal – 649–1078 – Santos et al. (2009a)
Rainwater (cold season), town of Aveiro Portugal – 28–157 – Santos et al. (2009b)
Continental rainwater – – 161 – Willey et al. (2000)
Marine rainwater – – 23 – Willey et al. (2000)
Rivers, Tributaries and Channels: Asia
Ganges River, Bangladesh portion Bangladesh, India ~2500 (975,000) 108–317 – Ittekkot et al. (1985)
Brahmaputra River, Bangladesh Portion Bangladesh, India,
China~2900 (580,000) 83–500 – Safiullah et al. (1987)
Yellow River (Huang He), mainstream in Upper
RegionChina ~5400 (745,000) 95–503 – Mostofa KMG et al.,
(unpublished data)Table 2 (continued)
(continued)

43 Dissolved Organic Matter in Natural Waters
(continued)Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Yellow river (Huang He) China ~5400 (745,000) 167–333 – Cauwet and Mackenzie
(1993), Zhang et al.
(1999)
Yangtze River (or Cháng Jiang) China ~6300 (1,950,000) 142–400 – Cauwet and Mackenzie
(1993), Zhang (1996)
Xijiang River, South China China – 84–495 – Gao et al. (2002)
Nunming River, Guizhou Proviance, South-west
ChinaChina 118 (−) 82–463 – Mostofa et al. (2010), Yue
et al. (2006)
NenJiang River (or Nen River), Heilongjiang Provi-
ance and Inner Mongolia, North-east ChinaChina 1370 (−) 346–1286 – Mostofa KMG et al.,
(unpublished data)
The Second Song Hua Jiang River, Jilin and Zheji-
ang proviance, North-east ChinaChina 849 (73,000) 362–1352 – Mostofa KMG et al.,
(unpublished data)
LiaoHe River, Heibei, Jilin, Inner Mongolia, Jilin,
and Liaoning Proviances, North-east ChinaChina 1394 (201,600) 169–1048 – Mostofa KMG et al.,
(unpublished data)
DaLingHe River, Inner Mongolia, North-east
ChinaChina 435? (13,000) 90–283 – Mostofa KMG et al.,
(unpublished data)
Rivers, Lake Biwa watershed (n = 10) Japan 20–60 32–375 – Mostofa et al. (2005b)
Kurose River, Hiroshima Prefecture Japan 43 (250) 123–385 – Mostofa et al. (2005b)
Ohta River, Hiroshima Prefecture Japan 103 (1,700) 45–164 – Mostofa et al. (2005b)
Rivers, Lake Fuxian and Lake Xingyun basin (n = 9)24șN – 37–428 – Hayakawa et al. (2004)
Yinluan River, Tianjin China – 396 – Tao (1996)
Bang Nara River; Saiburi River, Thailand 1–10șN – 125–625 – Yoshioka et al. (2002b)
Kahayan River and Rungan River, Indonesia 10șN–10șS – 558–1042 – Ishikawa et al. (2006)
Lanyang His River Taiwan – 42–667 – Kao and Liu (1997)
3 Rivers, Lake Hovsgol basin 50–52șN – 400–1400 – Hayakawa et al. (2003)
5 Major Rivers, Lake Baikal basin Russia – 43–542 – Yoshioka et al. (2002a, 2007)Table 2 (continued)

44 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Kuji River & 3 Tributaries, North Kanto Japan – 92–175 – Nagao et al. (2003)
Tributaries (n = 8) 30–35șN – 122–271 – Mostofa et al. (2007a)
Dongjiang tributary, South China Sea China – 186–227 – He et al. (2010)
Tributaries of NenJiang River North-East China – 177–2429 – Mostofa KMG et al.,
(unpublished data)
Tributary of LiaoHe River (n = 2) North-East China – 631–640 – Mostofa KMG et al.,
(unpublished data)
Irrigation Channels, Japan (n = 19) 30–35șN 3–5 65–210 – Mostofa et al. (2007a)
Africa
Zaïre (or Congo) River (Angola, Burundi,
Cameroon, Congo, Gabon, Rwanda, Tanzania,
Zambia)Africa 4700 (40,00,000) 250–833 – Martins and Probst (1991)
Niger River (Guinea, Mali, Niger, Benin, Nigeria) Africa 4180 (11,25,000) 167–500 – Martins and Probst (1991)
Gambia River (Republic of Guinea, The Gambia) West Africa 1130 (42,000) 83–333 – Martins and Probst (1991)
Europe
Lena River, East Siberia Russia 4472 (2,500,000) 309–1042 – Lara et al. (1998)
Rivers (n = 3), Arctic Russia 66–71șN 142–3917 – Duff et al. (1999)
Mura River, Slovenia Slovenia 438 (13,824) 150–708 – Brodnjak-V onˇ cina et al.
(2002)
River Tyne, UK 55șN (2,935) 275–1275 – Baker and Spencer (2004)
Rhône River Switzerland-
France(98,000) 94–818 – Rostan and Cellot (1995)
Scheldt River France, Belgium,
Netherlands~350 (21,600) 567 – Abril et al. (2000)
Rhine River (Germany, Italy, Austria, Switzerland,
France, Netherlands)– ~1233 (224,000) 242–442 – Eisma et al. (1982), Abril
et al. (2002)
Gironde River France, Spain ~602 (71,000) 258 – Veyssy (1998)Table 2 (continued)
(continued)

45 Dissolved Organic Matter in Natural Waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Thames River UK ~346 (14,000) 483 – Abril et al. (2002)
Elbe River Czech Republic,
Germany~1091 (145,800) 342–383 – Elbe (1997), Abril et al.
(2002)
Ems River Germany, Nether –
lands~371 (9,000) 567–675 – Cadée (1987), Abril et al.
(2002)
Sado River Portugal ~175 (7,600) 558 – Abril et al. (2002)
Douro River Portugal, Spain ~897 (115,320) 208 – Abril et al. (2002)
Loire River France ~1012 (115,000) 325 – Meybeck et al. (1988), Abril
et al. (2002)
Tagliamento River, Italy 46șN ~178 (2,900) 32–95 – Kaiser et al. (2004)
Vistula River (Poland, Ukraine, Belarus, Slovakia) – ~1000 (194,000) 397–653 – Guéguen et al. (2002)
Fuirosos, downstream rivers (summer season) Spain – 417–2750 – Vazquez et al. (2011)
Ebro River, NE Iberian Peninsula Spain ~900 (80, 000) 183–195 – Romani et al. (2004)
Öre River catchement, boreal watershed,
in northern SwedenSweden (2,940) 600–1558 – Bertilsson et al. (1999)
Latin America
Upstreams and Rivers Brazil 242–7500 – Daniel et al. (2002)
Paraná River, Argentina Argentina, Brazil,
Paraguay~4800 (2,582,000) 108–458 – Depetris and Kempe (1993)
Amazon River, Brazil Brazil, Colombia,
Ecuador, Peru~6800 (6,300,000) 235–1000 – Raymond and Bauer
(2001b), Richey et al.
(1990)
North America
Hudson River (New York, and New Jersey) USA 507 (21,000) 196 – Raymond and Bauer
(2001b)
York River (Virginia) USA 55 (4350) 390–701 – Raymond and Bauer
(2001b)
(continued)Table 2 (continued)

46 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
New River (North Carolina, Virginia and West
Virginia)USA 515 83–4167 – Newbern et al. (1981)
Parker River (Massachusetts) USA ~37 (609) 986 – Raymond and Bauer
(2001b)
Suwannee River (Florida and Georgia) USA 396 775–2583 – Crandall et al. (1999),
Schwede-Thomas et al.
(2005)
Ogeechee River, Georgia USA ~470 (14,000) 267–525 – See and Bronk (2005)
Almaha River, Georgia USA ~220 (36,000) 300–475 – See and Bronk (2005)
Satilla River, Georgia USA – 525–3133 – See and Bronk (2005)
Savannah River, Geogia USA – 225–417 – See and Bronk (2005)
Sapelo Island River, Geogia USA – ~50–2333 – Alberts and Takács (1999),
Haines (1979)
Satilla River and Altamaha River, Georgia USA ~378; ~220 1545–1620;
1430–1442– Xie et al. (2004)
Columbia River (Washington, D. C., Oregon and
British Columbia)USA, Canada 2000 117–158 – Prahl et al. (1998)
4 Major Rivers (Muskegon, Grand, St. Joseph and
Kalamazoo), Lake Michigan basinUSA – 430–1870 – Biddanda and Cotner
(2002)
Taylor River (Chennel), southern Everglades
National Park, FloridaUSA – 708–1533 – Davis et al. (2001)
Rivers (n = 8), southeastern United States USA – 417–2333 – Alberts and Takács (1999)
Hudson River & Mohawk River (New York) USA 507; 225 375–508 – Findlay et al. (1991)
Potomac River (Virginia, Maryland, Washington,
D. C.)USA 616 364 – Spiker (1981)
Susquehanna River, Maryland, Pennsylvania, New
YorkUSA 715 292 – Spiker (1981)Table 2 (continued)
(continued)

47 Dissolved Organic Matter in Natural Waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Rappahannock River, Virginia USA 314 125 – Spiker (1981)
James River, Virginia USA 560 167 – Spiker (1981)
Missouri River; Ohio River USA ~3700 (1,371,000);
1570 (490,000)283–308 – McKnight et al. (2001)
Mississippi River; and Atchafalaya River 28–29șN ~3734 (3,267,000);
220 (7,600)275–446 – Wang et al. (2004)
Delaware, Eel, and Mississippi River USA ~480 (36,500); 300
(9,500); -214, 80, 181 – Repeta et al. (2002)
3 Rivers (Shark, Broad and Caloosahatchee) 25–26șN – 170–1178 – Zanardi-Lamardo et al.
(2004)
2 Rivers (Flathead and Kootenai), Montana (n = 2) Canada, USA ~250 (22,700); 780
(50,000)40–667 – Perry and Perry (1991)
Yukon River Canada, Alaska ~3185 (800,000) 508–2835 – Guéguen et al. (2006)
Rivers, South-central Ontario Canada – 258–900 – Wu et al. (2005)
Mackenzie River, Arctic region Canada ~1700 (1,810) 312–576 – Osburn et al. (2009)
Lakes, Wetlands and Swamps: Asia
Lake Biwa, Japan 30–35șN (3,174) 88–183 (2.5–20 m) 76–101
(40–80 m)Mostofa et al. (2005a),
Sugiyama et al. (2004)
Lake Ashino, Japan (A1 and A2) Japan – 99–111 (0–10 m) 74–84 (30–38 m) Sugiyama et al. (2004)
Lake Ikeda, Japan ((I1 and I2) Japan – 101–112 (0–10 m) 55–56 (200–
233 m)Sugiyama et al. (2004)
Lake Suwa, Japan (Center) Japan – 142–216 (0 m) – Sugiyama et al. (2004)
Lake Inawashiro, Japan Japan – 42–47 (0–10 m) 26 (70 m) Sugiyama et al. (2004)
Lake Hongfeng & Lake Baihua, China 26șN – 169–330 (0–8 m) 161–194
(20–25 m)Fu et al. (2010)
Lakes, China (n = 3) 24–25șN – 98–614 105–123
(30–50 m)Sugiyama et al. (2004)
(continued)Table 2 (continued)

48 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Lake Ponds, (WangHua and TLH), North-east
China (n = 2)China – 2075–3152 – Mostofa KMG et al.,
(unpublished data)
Lake Fuxian, China 24șN – 116 95–100
(50–140 m)Hayakawa et al. (2004)
Lake Xingyun, China 24șN – 629–658 – Hayakawa et al. (2004)
8 Lakes (Batu, Tehang, Bunter, Bajawak, Pahewan,
Hampapak, Rengas & Takapan)Indonesia (10șN–
10șS)– 458–1300 – Ishikawa et al. (2006)
Lake Dapur and Lake Hurung) Indonesia (10șN–
10șS)– 2042, 4133 – Ishikawa et al. (2006)
Lake Hovsgol, Middle site, Mongolia 50–52șN – 85–100 85–90 (50–
200 m)Hayakawa et al. (2003)
Lake shore site toward middle of Lake Hövsgöl,
Mongolia50–52șN – 93–550 – Hayakawa et al. (2003)
Lake Baikal (North, Central & South basin) 51–55șN (556,000) 92–142 (0–100 m) 88–105 (500–
1620 m)Yoshioka et al. (2002a,
2007), Sugiyama et al.
(2004)
Lakes, New Zealand (n = 11) New Zealand (3.4–352) 25–833 – Rae et al. (2001)
Swamps, Bang Nara River basin, Thailand 1–10șN – 475–10333 – Yoshioka et al. (2002b)
Middle East
Lake Kinnerret Israel – 270–485 (0–10 m) 258–368 (38 m) Annual Report (2004)
Latin America
12 Lakes, Barciloche region Argentina – 21–222 – Morris et al. (1995)
Lake Barata, Brazil 1șN – 420–710 – Farjalla et al. (2006)
Europe – –
Lake Great Dun Fell UK – 333–3750 – Worrall et al. (2004a)
Lake Geneva, Switzerland–France Switzerland–
France– 59–128 – Guéguen et al. (2002)Table 2 (continued)
(continued)

49 Dissolved Organic Matter in Natural Waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Lakes (n = 38), Sweden 55–71șN – 167–1833 – Bertilsson and Tranvik
(2000)
Lakes, The United Kingdom Acid Waters Monitor –
ing Network (n = 11)UK (0.42–14.7) 67–550 – Monteith and Evans (2005),
Evans et al. (2006)
Lakes, Sweden (n = 5) Sweden – 325–1617
(1.5–7.5 m)– Granéli et al. (1996)
Lakes (n = 1000), Norway Norway – 167 (median) – Larsen et al. (2011)
North America
Lake Michigan USA – 124–216 – Biddanda and Cotner
(2002)
Lake Superior USA – 110–119 – Biddanda et al. (2001)
Lakes (Christmas, Turtle, Minnetonka, Owasso,
and Round)USA – 537–712 – Biddanda et al. (2001)
Lakes (Josephine, Johanna, Eagle, Medicine, and
Mitchell)USA – 545–785 – Biddanda et al. (2001)
Pony Lake and Lake Fryxel USA – 1789–4908 – Schwede-Thomas et al.
(2005)
Lakes, Michigan (n = 20) 46șN – 120–1764 – Pace and Cole (2002)
Banks Lake, and Okefenokee Swamp 25–30șN – 950, 2425 – Alberts et al. (1984)
Laramie River; Chimney Park wetland, USA 41șN – 592, 1250 – Brooks et al. (2007)
Lakes, Northeast USA, Colorado, Alaska (n = 47) USA – 67–1958 – Morris et al. (1995)
Lake Superior and Nobska Pond 47, 41șN – 117, 550 – Repeta et al. (2002)
Upper Great Lakes, North American temperate
forest (n = 188)USA – 133–2417 – Xenopoulos et al. (2003)
Seepage Lakes, central Maine (n = 17) USA – 132–1492 – David and Vance (1991)
Lakes (Sky Pond and The Loch), Loch Vale Water –
shed, ColoradoUSA – 67–308 – Baron et al. (1991)
(continued)Table 2 (continued)

50 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Redberry Lake and its three inflows, Canada 52șN – 933–2917 – Waiser and Robarts (2000)
Lakes (n = 59), southwestern Québec, Canada 51șN–74șW – 58–2350 – Houle et al. (1995)
Lakes (n = 20), Québec, Canada 45șN – 208–900 – Cammack et al. (2004)
Lakes: Saskatchewan lakes (n = 26), Canada 51–53șN – 342–13017 – Arts et al. (2000)
Ponds and 8 Wetlands (n = 10), Canada 52șN – 1283–6675 – Arts et al. (2000)
Prairie wetlands, Saskatchewan, Canada 52șN – 2933–10000 – Waiser and Robarts (2004)
Lakes, South-central Ontario (n = 4) Canada – 225–675 – Wu et al. (2005)
Freshwater Lakes (n = 23), Alberta Canada – 1833–5833 – Curtis and Prepas (1993),
Curtis and Adams (1995)
Saline Lakes (n = 35), Alberta Canada – 2667–27500 – Curtis and Prepas (1993)
Saline Lakes (n = 37), Prairies Canada – 4983–10000
(median)– Molot et al. (2004)
Non-saline Lakes (n = 32), Prairies Canada – 3250–3833
(median)– Molot et al. (2004)
Lakes (n = 45 & 11), Atlantic Maritime, Nova
Scotia and New BrunswickCanada – 483, 283 (median) – Molot et al. (2004)
Lakes (n = 60), Pacific Maritime Canada – 267 (median) – Molot et al. (2004)
Boreal lakes (n = 9), Ontario Canada – 150–425 – Hudson et al. (2003)
Lakes (n = 30), Boreal Plains Canada – 3000 (mean) – Molot et al. (2004)
Lakes (n = 42 & 16), Boreal Shield, Quebec and
NewfoundlandCanada – 508 (median) – Molot et al. (2004)
Lakes (n = 42 & 16), Boreal Shield, Quebec and
NewfoundlandCanada – 725, 358 (median) – Molot et al. (2004)
Lakes (n = 24), Boreal Cordillera Canada – 1225 (median) – Molot et al. (2004)
Lakes (n = 49), Taiga Shield Canada – 475–2042
(median)– Molot et al. (2004)Table 2 (continued)
(continued)

51 Dissolved Organic Matter in Natural Waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Lakes (n = 177), Montane Cordillera Canada – 125–1025
(median)– Molot et al. (2004)
Lakes (n = 38), Hudson Plains Canada – 783–1083
(median)– Molot et al. (2004)
Lakes (n = 37 & 18), Southern Arctic Canada – 525, 192 (median) – Molot et al. (2004)
Lakes (n = 12 & 6), Axel Heiberg and Victoria
Islands, Southern ArcticCanada – 192, 92 (median) – Molot et al. (2004)
Lakes (n = 9 & 1), Banks and Prince Patrick
Islands, Southern ArcticCanada – 283, 292 (median) – Molot et al. (2004)
Precambrian Shield lakes Canada – 357–1142 – Curtis and Schindler (1997)
Lakes (n = 12), Northwest Ontario, Canada 51șN (0.024–80000) 149–816 – Kelly et al. (2001)
Lakes (n = 3), Loch Vale Watershed, Rocky Moun-
tain National ParkCanada (204–660) 31–117 – McKnight et al. (1997)
Estuaries
Pearl River Estuary, upper to lower reaches China – 84–473 – He et al. (2010), Chen et al.
(2004)
Caeté Estuary, mangroves forest Brazil – 283–558 – Dittmar and Lara (2001)
Satilla Estuary, Geogia USA – 1972–2046 – Moran et al. (2000)
Hudson River Esturay USA – 217–539 – Hummel and Findlay
(2006)
York River Estuary USA – 254–713 – Raymond and Bauer
(2001a)
Coastal Estuaries (n = 2), San Francisco Bay and
Chesapeake Bay, USA37–39șN – 193 ± 4–222 ± 3 – Boyd and Osburn (2004)
Arctic Estuary, the Beaufort Sea Canada – 133–453 – Osburn et al. (2009)
Two Esturies, UK 55șN – 427–1427 – Baker and Spencer (2004)
(continued)Table 2 (continued)

52 K. M. G. Mostofa et al.Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Estuary, Denmark 55șN – 149–426 – Stedmon et al. (2003)
Adour Estuary, France France – 76–316 – Guéguen et al. (2002)
Gironde Estuary, French Atlantic coast France – 98–542 – Huguet et al. (2010)
Seine Estuary, French Atlantic coast France – 133–554 – Huguet et al. (2010)
Scheldt Estuary Europe – 183–517 – Abril et al. (2002)
Rhine Estuary Europe – 142–258 – Abril et al. (2002)
Gironde Estuary Europe – 92–208 – Abril et al. (2002)
Thames Estuary Europe – 217–417 – Abril et al. (2002)
Elbe Estuary Europe – 258–367 – Abril et al. (2002)
Ems Estuary Europe – 425–592 – Abril et al. (2002)
Sado Estuary Europe – 300–525 – Abril et al. (2002)
Douro Estuary Europe – 158–208 – Abril et al. (2002)
Loire Estuary Europe – 200–292 – Abril et al. (2002)
Coastal and open Oceans
Hiroshima Bay; Seto Inland Sea, Japan 32–34șN – 83–135 (0–30 m) 71–92 (50–
300 m)Mostofa KMG et al.,
(unpublished data)
Barguzin Bay; Coastal areas, Lake Baikal 51–55șN – 105–363 – Yoshioka et al. (2002a,
2007)
Mississippi River Plume 28–29șN – 54–124 – Chen et al. (2004)
Orinoco River Plume USA – 70–276 – Blough et al. (1993), del
Castillo et al. (1999)
Chesapeake Bay 35–41șN – 118–215 – Mitra et al. (2000)
Florida Bay 24–25șN – 139–147 – Zanardi-Lamardo et al.
(2004)
West Florida Shelf USA – 89–305 – del Castillo et al. (2000)
Middle Atlantic Bay USA – 70–150 – V odacek et al. (1995)Table 2 (continued)
(continued)

53 Dissolved Organic Matter in Natural Waters
Study sites Country/ Length/ DOC Reference
latitudes watershed areaaSurface (Depth)cDeeper (Depth)b
km or (km2)a(μM C)
Kara Sea, an estuary of Arctic Ocean 72–73șN – 423–537 – Opsahl et al. (1999)
Middle Atlantic Bight (MAB) USA – 82–98 48–90 (90–
2600 m)Mitra et al. (2000)
Arctic shelf, the Beaufort Sea Canada – 97–229 – Osburn et al. (2009)
Arctic Gulf, the Beaufort Sea Canada – 70–126 – Osburn et al. (2009)
South Baltic Sea 53–66șN – 474–616 – Ferrari et al. (1996), Ferrari
and Dowell (1998)
Northern Gulf of Maxico 35–41șN – 50–100 – Wang et al. (2004)
North Pacific Ocean 31șN – 87 38–55 (200–
6000 m)Williams and Druffel (1987)
Western North Pacific 35șN – 85–117 66–73 (150 m) Ogawa and Ogura (1992)
Arctic Ocean 74–81șN – 58–85 (50–200 m) 49–54 (>1000 m) Opsahl et al. (1999)
Atlantic Ocean 3–75șN – 50–97 (<100 m) 33–59 (>1000 m) Ogawa and Tanoue (2003)
Pacific Ocean 0–58șN, 1267șS – 40–90 34–45 Ogawa and Tanoue (2003)
Indian Ocean; Arabian Sea 5–20șN – 55–95 43 Ogawa and Tanoue (2003)
Antarctis Ocean 42–77șS – 38–75 34–60 Ogawa and Tanoue (2003)
Arctic Ocean 70–85șN – 34–107 49–54 Ogawa and Tanoue (2003)
“n” indicates the number of upstreams, rivers and lakes whereas DOC levels are mentioned for ranges or average of the samples
ameans the watershed area (km2) applicable to lakes and the values in parentheses are river watershed area
bmeans the water depth in lakes and oceans and values are mentioned within the bracket
cmeans the surface layer and depth belongs within 0–50 mTable 2 (continued)

54 K. M. G. Mostofa et al.
newly flooded terrestrial vegetation, and leaching of organic matter from flood-
plain soils. In Latin American Rivers, the major sources of DOC are the dumping of untreated urban sewage into the rivers, the leaching of DOM from humid ever –
green and deciduous forests, floodplains of ox-bows and ponds, seasonally humid savannas and shrublands, as well as swamps. The sources of DOC in European Rivers may be natural such as wetlands, swamps and peat-land or anthropogenic such as industrial activities and sewerage. The major sources in North America that account for the high DOC levels are natural (wetlands, swamps, marshlands, ground water of mountainous areas covered by riparian vegetation ecosystems) and anthropogenic (agricultural, sewerage and industrial effluents).
In groundwater, DOC concentrations are 16–424 μM C in Asia, 183–842 μM
C in Australia, 42–15333 μM C in Europe, 8–8333 μM C in North America, 100–3000 in Brazil, 1108 ± 217–14167 ± 6333 μM C in Botswana, and finally 167–7500 μM C in soil solution in North America and Europe (Table 2) (Thurman
1985a; Mostofa et al. 2007a, Mostofa KMG et al., unpublished data; Buckau et al. 2000; Bertilsson et al. 1999; McIntyre et al. 2005; Mladenov et al. 2008; Meier
et al. 2004; Gielen et al. 2011; Crandall et al. 1999; Schwede-Thomas et al. 2005;
Pabich et al. 2001; Michalzik et al. 2001; Anawar et al. 2002; Richey et al. 2002;
Bradley et al. 2007). The results of the available studies show that DOC concen-trations in groundwater are significantly higher in Europe, North America, South America, and African regions than in Asia. High ground-water DOC concentra-tions are directly related to the high DOC concentrations in surface waters and high infiltration rates (Mladenov et al. 2007).
In rainwater, DOC concentrations are substantially varied among different coun-
tries, such as 25–3675 μ M C in China, 217 μ M C in Brazil, 4–1908 μ M C in USA,
28–1078 in Europe, 161 μ M C in a continental location, and 23 μ M C in a marine
location (Table 2 ) (Likens et al. 1983; Guggenberger and Zech 1993; Sakugawa
et al. 1993; Willey et al. 2000; Ciglasch et al. 2004; Avery et al. 2006; Kieber
et al. 2006; Santos et al. 2009a, b; Southwell et al. 2010; Miller et al. 2008;
Mostofa KMG et al., unpublished data; Xu et al. 2008; Pan et al. 2010). Rainwater DOC concentrations are significantly high in China and Europe, but relatively low in USA, and lower in marine locations than in continental ones. These results sug-gest that DOC concentrations in rainwater are mostly affected by the atmospheric organic contaminants, which generally derive from the surrounding environments.
DOM Contents in Lakes, Wetland and Swamps
DOC concentrations are significantly variable in lakes situated among differ –
ent continents: 80–4133 μ M C in Asia (Table 2 ) (Yoshioka et al. 2002a, 2007; Fu
et al. 2010; Mostofa et al. 2005a, Mostofa KMG et al., unpublished data;
Hayakawa et al. 2003, 2004; Sugiyama et al. 2004; Ishikawa et al. 2006; Rae et al.
2001); approximately 21–710 μ M C in Latin America (Morris et al. 1995; Farjalla
et al. 2006); approximately 59–3750 μ M C in Europe (Bertilsson and Tranvik
2000; Lu et al. 2007; Worrall et al. 2004a; Evans et al. 2006; Monteith and Evans
2005; Guéguen et al. 2002; Larsen et al. 2011); and approximately 70–27500 μ M

55 Dissolved Organic Matter in Natural Waters
C in North America (Morris et al. 1995; Pace and Cole 2002; Hudson et al. 2003;
Wu et al. 2005; Cammack et al. 2004; McKnight et al. 1997; Allard et al. 1994;
Waiser and Robarts 2000; Biddanda and Cotner 2002; Repeta et al. 2002; Schwede-Thomas et al. 2005; Curtis and Prepas 1993; Curtis and Schindler 1997; Arts et al. 2000; Biddanda et al. 2001; Kelly et al. 2001; Xenopoulos et al. 2003; Molot et al. 2004; Brooks et al. 2007). The general results of DOC concentrations studies
in lakes show some characteristic phenomena (Table 2 ): (i) relatively low DOC
concentrations have been found in Asia, but lakes situated Indonesia and North-East China shows high contents of DOC. (ii) Lakes situated in Europe and North America show relatively high DOC values. DOC concentrations in Saline lakes are significantly higher than in non-Saline lakes situated in the Prairies and Alberta in Canada. Boreal lakes often show high contents of DOC. This is interesting because the boreal region contains roughly 30 % of the global lakes and their water is rich in OM (Molot and Dillon 1996; Downing et al. 2006; Benoy et al. 2007).
A large lake database (7,514 lakes from 6 continents) shows that mean DOC
concentrations are 632 ± 16 μM C, ranging from 8 μM C to 27667 μM C (Sobek
et al. 2007). In 87 % of the lakes DOC was between 83 and 1667 μ M C, whilst
8.3 % of the lakes had concentrations lower than 83 μ M C. Lakes between 1667
and 3333 μ M C are relatively few (4.2 %), and only 0.4 % of the lakes had DOC
concentrations above 3333 μM C (Fig. 2a). In 55 % of the lakes DOC concentra-
tion was above 417 μ M C, which is suggested to be a threshold value for the tran-
sition between net autotrophy and net heterotrophy in lakes (Sobek et al. 2007; Jansson et al. 2000; Prairie et al. 2002). Several hypotheses are considered based on DOC concentration and terrestrial vegetation (Sobek et al. 2007). First, arctic lakes are generally characterized by low DOC concentrations where land-cover types are wooded tundra to bare desert and correspond to an annual mean temper –
ature of <− 4 °C (Fig. 2 b). Second, boreal lakes display a tendency toward higher
DOC concentration if the land-cover types are deciduous and mixed boreal forest to deciduous conifer forest, which corresponds to an annual mean temperature of 0.5– 4 °C (Fig. 3 b). Third, lakes on the northern Great Plains in Saskatchewan
and Canada show the highest DOC concentrations (~80 to ~10500 μ M C) and
the land-cover types are cool grasses, shrubs, cool crops and towns (Fig. 3c).
Fourth, Fig. 2 b shows that DOC concentrations for lakes situated in the warmer
climate zones (average mean annual temperature >17.4 and <22.3 °C) reach rela-tively high values (approximately 80–3300 μ M C) for land-cover types including
broadleaf crops, corn and cropland and conifer forests (Fig. 2 b). Correspondingly,
DOC concentrations for lakes situated in the low warmer climate zones (average mean annual temperature >8.0 and <13.3 °C) are relatively low (approximately 50–2000 μM C). Here typical land-cover types are crops, mixed and deciduous
broadleaf forests, cool and cool rain forests, grass, and shrubs. Fifth, Fig. 2 b sug-
gests that DOC concentrations for lakes that experience relatively low annual temperature (>1.5 and <4.3 °C) are relatively high (approximately 50–2100 μ M
C), and the land-cover types are mostly woods, deciduous and mixed boreal for –
est, cool mixed forest, conifer boreal forest, narrow conifers and deciduous coni-fer forests.

56 K. M. G. Mostofa et al.
DOM Contents in Estuaries
DOC concentrations in estuaries are often lower than in lakes all over the world.
They reach 84–473 μM C in China (Chen and Gardner 2004; He et al. 2010); 283–558 μM C in Brazil (Dittmar and Lara 2001); 190–2046 μM C in the USA
(b)
(c)(a)
Fig. 2 Box and whisker-plot of the distribution of DOC concentration (a) for all lakes, (b) for
lakes divided into different land-cover types, and (c) for lakes in Saskatchewan, Canada. In (b),
the land-cover types have been sorted according to the average mean annual temperature. The boxes display the median and the quartiles, the whiskers represent the 10 and 90 % percentiles, and the points represent the 5 and 95 % percentiles. Only land-cover types containing 10 or more lakes are shown. The land-cover types “inland water” and “sea water” are omitted from the plot because they are not indicative of climate or geography. Data source Sobek et al. (2007)

57 Dissolved Organic Matter in Natural Waters
(Moran et al. 2000; Raymond and Bauer 2001a; Boyd and Osburn 2004; Hummel
and Findlay 2006); 133–453 μM C in Canada (Osburn et al. 2009); 76–1427 μM
C in Europe (Abril et al. 2002; Guéguen et al. 2002; Huguet et al. 2010; Baker and Spencer 2004; Stedmon et al. 2003). DOC in estuaries is originated from both autochthonous sources (algae or phytoplankton) and allochthonous sources such as material of terrestrial plant origin.
DOM Contents in Coastal and Open Oceans
DOC concentrations are substantially higher (50–616 μM C) in coastal seawa-
ters than in the open ocean (40–117 μM C) (Table 2) (Ogawa and Tanoue 2003; Mitra et al. 2000; Chen and Gardner 2004; Williams and Druffel 1987; Opsahl and Benner 1998; Wang et al. 2004; Zanardi-Lamardo et al. 2004; Osburn et al. 2009;
Blough et al. 1993; V odacek et al. 1995; Ferrari et al. 1996; Ferrari and Dowell 1998; Del Castillo et al. 1999, 2000; Mostofa KMG et al., unpublished data).
DOC concentrations in coastal waters are generally regulated by terrestrial or riv-erine input, zooplankton feeding and algal or phytoplankton production (Lee and Wakeham 1988; Lee and Henrichs 1993; Mann and Wetzel 1995; Hedges et al.
1997). Increased biomass of primary producers also plays a major role in regulat-ing the DOM contents in coastal areas (Ittekkot 1982; Billen and Fontigny 1987;
Bronk et al. 1998). DOC concentrations in open oceans are relatively low and are included in the range of 40–117 μM C at epilimnion and 30–90 μM C at hypolim-nion (Table 2) (Ogawa and Tanoue 2003). DOC values appear to be relatively uni-form in most of the oceans, whilst in coastal areas the DOM pool becomes much more heterogeneous because of terrestrial inputs. In open oceans, the DOC con-centrations in epilimnetic layers follow the order of Arctic Ocean (70–107 μM C) > subtropical zone (~80 μM C) > tropical (equatorial) and temperate zones (60–70 μM C) > subarctic and subantarctic regions (50–60 μM C) > Antarctic
region (40–60 μM C) (Ogawa and Tanoue 2003). There are two main sources of DOC in coastal oceans: allochthonous DOM of terrestrial origin and autochtho-nous DOM of algal or phytoplankton origin. In open oceans, allochthonous DOM contents are gradually decreased photolytically depending on the distance from coastal areas, whilst autochthonous DOM is substantially increased (see the con-tribution of DOM for detailed discussion). The three major sources of DOC in the Arctic Ocean are in situ production (56 %), river run-off (25 %), and Pacific water (19 %) (Kirchman et al. 1995).
6 Factors Affecting DOM in Natural Waters
DOM contents and its dynamics are mostly dependent on two issues: origin and/or input, as well as its consequent mineralization by various environmental fac-tors that are associated with the watershed activities of natural waters. The river-groundwater interface can act as a source or sink for DOM, depending on the volume and direction of flow, DOC concentrations and biotic activity (Brunke

58 K. M. G. Mostofa et al.
and Gonser 1997). The dynamics of lake DOM is greatly affected by specific
and regional factors such as pH of lake water, air temperature, solar radiation, precipitation, sulfate deposition, DOC contents in the adjacent rivers, vegeta-tion of the terrestrial ecosystem, and southern oscillation index (SOI) (Mostofa
et al. 2005a, 2009b; Hudson et al. 2003). A complex balance of abiotic and biotic
processes controls the molecular composition of marine DOM to produce signa-tures that are characteristic of different environments (Kujawinski et al. 2009). Therefore, the controlling factors affecting the origin and dynamics of both allochthonous and autochthonous DOM in natural waters can be distinguished as:
(i) Types and nature of terrestrial plant material in soil; (ii) Land management and natural effects (precipitation, flood and drought); (iii) Effect of temperature;
(iv) Microbial processes; (v) Photoinduced processes; (vi) Photosynthesis in natural waters; (vii) Metal ions complexation and salinity; and (viii) Global warming.
6.1 Types and Nature of Terrestrial Plant Material in Soil
Allochthonous DOM is originated in soil by microbial degradation of leachable organic carbon, which varies depending on the types and nature of terrestrial plant communities, soil types and other regional effects (Mostofa et al. 2009a; Nakane et al. 1997; Uchida et al. 1998, 2000; Moore et al. 2008; Tu et al. 2011; Kindler
et al. 2011; Duff et al. 1999; Michalzik et al. 2001; Rae et al. 2001; Cronan and
Aiken 1985; Frost et al. 2006; Johnson et al. 2006). The litter-rich surface soils
have relatively higher DOC concentration than the litter-lacking ones, which can be distinguished because in the former case the δ
13C values of DOC are closer to
the δ13C of litter than to the δ13C of organic carbon in forest soil (Tu et al. 2011).
In most temperate and boreal landscapes the DOC concentrations in inland waters are regulated by a wide variety of watershed characteristics, including the quan-tity and type of vegetation, watershed slope, and particularly the extent and nature of wetlands (Kindler et al. 2011; Allard et al. 1994; Rae et al. 2001; Xenopoulos
et al. 2003; Frost et al. 2006; Engstrom 1987; Williamson et al. 2001; Rice 2002;
Canham et al. 2004; Winn et al. 2009).
DOC leaching from topsoils in the presence of different vegetation is largely
variable. Therefore, different values have been observed in the presence of grass-lands (range 158–1425 μM C and mean: 667 μM C), croplands (range: 325–
1442 μM C and mean: 1000 μM C) and forests (range: 592–3592 μM C and
mean: 1917 μM C), but large variations have also been observed within land use classes (Kindler et al. 2011). Under Cerrado vegetation, total organic C (TOC)
concentrations (filtered < ca. 1 μm) found in the soil solution (ca. 417 μM C) between 15 and 200 cm depth were lower than those usually found in the soil of temperate forests (833–1667 μM C) (Michalzik et al. 2001; Lilienfein et al. 2001). TOC concentrations in the soil solution under Pinus are lower than under Cerrado (Lilienfein et al. 2001). In uplands, soils derived from coniferous forests

59 Dissolved Organic Matter in Natural Waters
are much richer in DOC than those from hardwood stands or grass lands (Rae
et al. 2001; Cronan and Aiken 1985). Lakes with catchments containing 55–65 %
natural grassland and <30 % forest have low DOC concentration (<83 μM C), whilst lakes in moderately forested (50–60 %) catchments have DOC concentra-
tions of 83–208 μM C, and those in densely forested (>70 %) catchments have DOC concentrations of 667 μM C (Cronan and Aiken 1985). Forested wetlands,
particularly those with coniferous trees, are positively related to lake DOC, open-water wetlands including lakes are inversely related to DOC, and scrub-shrub and emergent wetlands are not related to DOC (Xenopoulos et al. 2003). Upstream rivers covered by coniferous, deciduous or moxed-type forests have generally low (~<200 μM C) DOC concentrations (Table 2) (Mostofa et al. 2005a, b; Sugiyama
et al. 2005). A model analysis has shown that the terrestrial land cover such as conifer boreal forest and barren tundra strongly affects DOC in lakes (Sobek et al. 2007). The land cover type “conifer boreal forest” is positively related with lake DOC, while “cool conifer forest” is negatively related to DOC. However, cool conifer forest is confined to high altitude areas such as the Rocky Mountains and the Alps, which may explain the relatively low DOC concentrations found in these lakes (Sobek et al. 2007).
In polar desert lakes, DOM is generated autochthonously by microbial pro-
cesses in water, since there is no catchment vegetation (Rae et al. 2001). This DOM thus differs from that of temperate latitudes by having a reduced ratio of aromatic to aliphatic residues (McKnight et al. 1994). It has also been shown that DOC from predominantly grassland catchments is qualitatively different in terms of its UVR attenuation properties than DOC from a mainly forested catchment (Rae et al. 2001). Therefore, the type and amount of terrestrial vegetation sur –
rounding a catchment plays a significant role in defining the concentration levels of DOC in the catchment water. It is hypothesized that high ground-water DOC concentrations are directly related to high DOC concentrations in surface waters (Mladenov et al. 2007).
6.2 Land Management and Natural Effects (Precipitation,
Flood and Drought)
Land management and natural effects (precipitation, flood and drought) are
important factors for controlling DOM release from soil environments to natural water catchments (Mostofa et al. 2005b, 2007a; Watts et al. 2001; Ittekkot et al.
1985; Safiullah et al. 1987; Newbern et al. 1981; Richey et al. 1990; Depetris and Kempe 1993; Shaw 1979; Worrall et al. 2003; Worrall and Burt 2004; Yallop and
Clutterbuck 2009; Clutterbuck and Yallop 2010; Yallop et al. 2010). These pro-
cesses include several phenomena:
(i) DOC is largely released from soil into water during agricultural activities,
particularly in plantation and growing seasons of rice plants as well as other

60 K. M. G. Mostofa et al.
plantations through rainfall or water overflow (Mostofa et al. 2005a, 2007a).
Rapid photo- and microbial respiration or assimilation of soil OM might be
responsible for high releases of DOC into water during agricultural activities.
(ii) DOC concentrations are increased with afforestation in catchments (Ciglasch et al. 2004; Neal et al. 1998, 2004). In addition, deforestation can reduce
evaporation and increase surface temperature. Changes in land-surface cover can enhance the degradation of soil DOM and OM by both photoinduced and microbial processes (Brandt et al. 2009; Rutledge et al. 2010; Raich and
Schlesinger 1992; Borges et al. 2008). Therefore, either afforestation or defor –
estation in soil environments can contribute to the rapid washout of allochtho-nous DOM and OM to water catchments by precipitation or runoff.
(iii) Controlled heather burning as a management tool for red grouse (Lagopus lagopus) husbandry in peat surface is often (but not always) identified as a highly significant driver of spatial variance in DOC concentration in drain-age water (Yallop and Clutterbuck 2009; Yallop et al. 2006, 2008; Ward et al.
2007; Worrall et al. 2007).
(iv) Increased precipitation and runoff can lead to higher DOC export from the catchment into natural surface waters (Pace and Cole 2002; Zhang et al. 2010;
Monteith et al. 2007; Hongve et al. 2004; Sobek et al. 2007; Ciglasch et al. 2004; Gielen et al. 2011; Anderson et al. 1997; Evans et al. 1999). Total solar radiation and precipitation can account for 49–84 % of the variation in the long-term DOC patterns in various catchments (Zhang et al. 2010). The DOC concentrations in Swedish lakes and streams have substantially increased dur –
ing the 1970–1980s, mostly due to higher precipitation (Tranvik and Jasson 2002). DOC concentrations vary from 4 μM C to 3675 μM C in rainwater, which may largely affect the natural surface waters (Table 2) (Likens et al. 1983; McDowell and Likens 1988; Guggenberger and Zech 1993; Chebbi and
Carlier 1996; Willey et al. 2000, 2006; Ciglasch et al. 2004; Avery et al. 2006;
Kieber et al. 2006, 2007; Miller et al. 2009; Santos et al. 2009a, b; Southwell
et al. 2010; Pan et al. 2010). Factors affecting the variation in DOC concen-trations are rainwater volume, season (winter, spring or summer), location of the rain events, wind speed, storm trajectory, and the air mass pathways dur –
ing precipitation. During highly rainy seasons, DOC concentrations are sig-nificantly increased through flushing of organic-rich waters from upper soil horizons, agricultural and forest runoff, primary production and subsequent flooding of the ox-bow or flood-plain lakes, through lake out-flowing into the nearby catchment waters (Mostofa et al. 2005b; Ittekkot et al. 1985; Safiullah et al. 1987; Ishikawa et al. 2006; Newbern et al. 1981; Richey et al. 1990; Depetris and Kempe 1993; Mulholland 2003). Leaching and heterotrophic processing of newly flooded terrestrial vegetation, leaching of organics from floodplain soil and catchment areas can also affect the DOC concentrations (Wetzel 1992; Duff et al. 1999; Mladenov et al. 2007; Reche et al. 1999). The
flow path of water in the catchment after precipitation is an important factor that can reduce soil erosion (Sobek et al. 2007), thereby reducing the rapid washout of allochthonous DOM, POM and nutrients to water catchments.

61 Dissolved Organic Matter in Natural Waters
Many studies observe that DOC export is positively correlated with runoff,
and two issues are involved (Sobek et al. 2007): First, the carbon budget of the studied landscapes is not in steady state, i.e., increased runoff exports more DOC than is produced in soil, which implies that observed increases are temporary. Second, changes in runoff are concomitant to changes in leachable organic carbon stocks in soil. High runoff indicates a high water table, which favors DOC leaching and hampers microbial degradation due to anoxia and humification. In addition, the negative relationship between runoff and lake DOC concentration indicates that when high runoff prevails over extended periods of time, the leachable soil organic carbon pool will eventually be reduced (Sobek et al. 2007).
(v) Severe drought seasons can either greatly decrease or increase the DOC lev-els in natural water (Watts et al. 2001; Meier et al. 2004; Shaw 1979; Worrall et al. 2003, 2005, 2006; Worrall and Burt 2004; Ward et al. 2007). DOC con-
centrations in shallow groundwater are very low (1558 μM C) under drought
condition compared to spring samples (2583 μ M C). Correspondingly, the
properties of DOM are largely different (Meier et al. 2004). Such differences may be attributed to biogeochemical changes in the DOM pool over the sum-mer and fall seasons under drought conditions. Studies show that runoff char –
acteristics and flow-paths within peat soils change as a result of severe drought, which could increase DOC concentrations in runoff water (Evans et al. 1999; Holden and Burt 2002).
In addition, releases of allochthonous DOM are largely dependent on several
catchment properties such as drainage ratio (catchment : lake area), proportion of wetlands, proportion of upstream lakes, watershed slope, altitude, catchment area, % peat cover, water area, wetland cover, and soil C:N ratio in the catchment (Sobek et al. 2007; Eckhardt and Moore 1990; Xenopoulos et al. 2003; Rasmussen et al. 1989; Kortelainen 1993; Hope et al. 1997; Aitkenhead and McDowell 2000).
The export of DOC from catchments is often related to the organic carbon stocks in the catchment soils (Hope et al. 1994; Aitkenhead et al. 1999). Studies show that DOC fluxes are small: 0.8 ± 0.2 % relative to gross primary productivity, 1.0 ± 0.3 % relative to ecosystem respiration, and (2.4 ± 0.4 %) relative to soil
respiration, when the DOC fluxes are considered relative to the gross ecosystem carbon fluxes in a specific catchment (Gielen et al. 2011).
6.3 Effect of Temperature
Temperature can affect DOM in two ways: First, water temperature (WT), linked with solar radiation, is one of the most important variables in the production of autochthonous DOM in natural waters because it affects the physical, pho-toinduced, microbial and ecological processes (Mostofa et al. 2009a; Sobek et al. 2007; Gielen et al. 2011; Gudasz et al. 2010). The mineralization of organic

62 K. M. G. Mostofa et al.
carbon in lake sediments exhibits a strong positive relationship with temperature,
which suggests that increasing temperature would lead to increased mineraliza-tion of OM in natural waters (Gudasz et al. 2010). As already seen, the distribution of DOC concentrations for various lakes (7500 lakes from 6 continents; Fig. 2 )
(Sobek et al. 2007) does not show a simple relationship between DOC and tem-perature, because the DOC values are affected by both temperature and the sur –
rounding vegetation. However, autochthonous DOM is often higher in lake waters where the water temperature (WT) is higher. For example, WT in the surface water of Lake Baikal is generally lower (4–16 °C: summer period) compared to Lake Biwa (10–28.7 °C: summer period), although the DOC levels are almost similar: 88–114 μM C at 0–1400 m depth and 76–135 μ M C at 0–80 m at central basins,
respectively, during the summer stratification period (Weiss et al. 1991; Mostofa
et al. 2005a; Yoshioka et al. 2002a; Goldman et al. 1996). However, autochtho-nous production in Lake Biwa is significantly higher (3–82 %) than in Lake Baikal (6–35 %) (Table 2 ) (Mostofa et al. 2005a; Yoshioka et al. 2002a; Sugiyama et al.
2004). Autochthonous production is not observed in a region where the WT is very low (ca. ≤ 0 °C) and, at the same time, chlorophyll a (Chl a ) production does not
occur in the upper water column (Bussmann and Kattner 2000) However, a little increase in WT may produce a little amount of Chl a with a corresponding increase
in autochthonous DOM in natural waters (Wheeler et al. 1996, 1997; Bussmann
and Kattner 2000; Melnikov and Pavlov 1978; Tremblay et al. 2006). Low WT
may affect the DOM contents by several pathways (Mostofa et al. 2009a): (i) Photoinduced degradation of surface DOM is less effective due to low solar effects at low WT and air temperature. This may result in low contents of photoprod-ucts, such as DIC, CO
2, H2O2, LMW organic substances and so on, which subse-
quently decreases photosynthesis and primary production. The result is a decrease of autochthonous DOM production in natural waters. (ii) Mineralization of DOM by photoinduced degradation becomes significantly low at low WT, which may preserve the DOM in natural waters and lead to increased allochthonous DOM contents.
An increase in air temperature can significantly enhance DOC export from soil
to surface water by increasing soil respiration and mineralization of plant organic material (Mostofa et al. 2005a, b; Monteith et al. 2007; Raymond and Saiers 2010;
Gielen et al. 2011; Evans et al. 1999, 2002; Gudasz et al. 2010; Newson et al.
2001). This can lead to DOM leaching from groundwater to stream or riverbeds or lakes, and the DOC concentrations are linearly increased with increasing tem-perature in natural waters. Coherently, the DOC leaching from forest catchments to streams is significantly enhanced during the summer season. It has been shown that releases of DOM in upstream waters of forest mountainous origin are much higher (28–84 %) during summer than in winter season. This has been estimated during monthly samplings in four upstream rivers and the releases were highest (52–84 %) in forest soils at upstream sites of Kurose River than in forest gran-ite mountain (28–31 %) in Lake Biwa watershed (Mostofa et al. 2005a, b). It is
suggested that increased temperature during the summer season can lead to higher microbial activity and enhanced decomposition of organic matter or peat, which

63 Dissolved Organic Matter in Natural Waters
increases production of DOC. Increases in temperature can also induce higher
drawdown of water tables in summer, increasing the depth of the zone where oxi-dation and production of DOC take place (Evans et al. 1999). In the UK, the latter effects would probably be exacerbated by the decreased summer rainfall over the last 40 years (Burt et al. 1998). It is suggested that the effect of increased tempera-ture on water tables can account for between 10 and 20 % of the increase in DOC concentration (Worrall et al. 2004b, 2007; Cole et al. 2002).
6.4 Microbial Processes
Microbial processes have two important effects on OM (DOM and POM). First microbial respiration or assimilation of OM into algal or phytoplankton bio-mass or bacterial biomass can release autochthonous DOM in deep water (Mostofa et al. 2009a, b, 2011; Zhang et al. 2009; Yamashita and Jaffé 2008; Fu et al. 2010;
Rochelle-Newall and Fisher 2002a; Yamashita and Tanoue 2004; Aoki et al. 2008;
Stedmon and Markager 2005a; Stedmon et al. 2007a). This process can give an important contribution to autochthonous DOM in natural waters. Second, microbial processes can change the molecular structure of DOM components and their opti-cal properties, either absorption properties of CDOM or fluorescence properties of FDOM (Moran et al. 2000; Mostofa et al. 2007a; Hur 2011). Such properties will
be discussed in chapters “Photoinduced and Microbial Degradation of Dissolved Organic Matter in Natural Waters”, “Colored and Chromophoric Dissolved Organic Matter in Natural Waters”, and “Fluorescent Dissolved Organic Matter in Natural Waters”. Microbial degradation can mineralize DOC by approximately 0–85 % in natural waters (see chapter “Colored and Chromophoric Dissolved Organic Matter in Natural Water”). High molecular weight protein-like structures in plant-derived DOM are degraded primarily through physical–chemical and microbial processes (Scully
et al. 2004). Microbial activity is significantly stimulated by the photoproducts of read-ily assimilable nitrogen compounds such as ammonium and amino acids (Bushaw
et al. 1996; Jørgensen et al. 1998). Under N-limiting conditions, nitrogenous photo-products can significantly increase the rates of bacterial growth in natural waters (Bushaw et al. 1996). Microbial degradation depends on several key factors, such as occurrence and nature of microbes; sources of DOM and the quantity of its fermenta-tion products; temperature; pH; and sediment depth (see chapter “Photoinduced and Microbial Degradation of Dissolved Organic Matter in Natural Waters”).
6.5 Photoinduced Processes
Photoinduced processes have two effects on OM (DOM and POM): First, photo-res-piration or assimilation of POM can release autochthonous DOM in surface waters (Mostofa et al. 2009b, 2011; Harvey et al. 1995; Fu et al. 2010; Rochelle-Newall

64 K. M. G. Mostofa et al.
and Fisher 2002a; Hiriart-Baer and Smiith 2005). This process can give a signifi-
cant contribution to autochthonous DOM in natural waters. Second, solar radiation causes changes in the molecular structure of DOM and decomposes its functional groups. This effect can be detected chemically as mineralization of DOC, by approximately 0–54 % during irradiation times ranging from hours to months. It can also be detected optically as alteration of either chromophoric dissolved organic matter (CDOM) or fluorescent dissolved organic matter (FDOM) (see chapters “Photoinduced and Microbial Degradation of Dissolved Organic Matter in Natural Waters”, “Colored and Chromophoric Dissolved Organic Matter in Natural Waters”, and “Fluorescent Dissolved Organic Matter in Natural Waters”). Photoinduced degradation can also reduce the mean molecular size of the high molecular weight DOM (Lovley and Chapelle 1995; Lovley et al. 1996; Yoshioka et al. 2007; Amador et al. 1989), which subsequently produces low molecular weight
(LMW) intermediates (Lovley et al. 1996; Wetzel et al. 1995; Amon and Benner 1994; Dahlén et al. 1996). This process ultimately ends up in mineralization with formation of e.g. COS, CO, CO
2, DIC, ammonium and gaseous hydrocarbons
(Miller and Zepp 1995; Miller 1998; Johannessen and Miller 2001; Ma and Green 2004; Xie et al. 2004; Johannessen et al. 2007; Gennings et al. 2001; Clark et al. 2004). Photoinduced degradation generally occurs in the mixing zone and decreases with an increase in water depth (Bertilsson and Tranvik 2000; Ma and Green 2004;
Vähätalo et al. 2000; Mostofa et al. 2005a; Granéli et al. 1996). The photoreactivity
of fluorescent DOM is greatly decreased when passing from freshwater to marine waters, but deep waters in lakes or marine environments are often more sensitive to photoinduced degradation processes than surface waters (Mostofa et al. 2011). Similar effects have been observed as far as photomineralization is concerned (Vione et al. 2009). Photoinduced degradation is significantly affected by several key factors, such as solar radiation, water temperature, effects of total dissolved Fe and photo-Fenton reaction, occurrence and quantity of NO
2− and NO 3− ions,
molecular nature of DOM, water pH and alkalinity, dissolved oxygen (O 2), water
depth, physical mixing in the surface mixing zone, increased UV-radiation during ozone hole events, global warming, and salinity (see chapter “Photoinduced and Microbial Degradation of Dissolved Organic Matter in Natural Waters”).
6.6 Photosynthesis in Natural Waters
Autochthonous production of DOM in natural waters is mostly accompanied by photosynthesis (Takahashi et al. 1995; Hamanaka et al. 2002; Marañòn et al. 2004). Photosynthetically produced POM (mostly algae or phytoplankton) and the related release of new DOM are significantly influenced by several key factors, such as high precipitation (Freeman et al. 2001a; Tranvik and Jasson 2002; Hejzlar et al. 2003; Zhang et al. 2010), nitrogen deposition (Pregitzer et al. 2004; Findlay 2005), sulfate (SO
42−) deposition (Zhang et al. 2010; Evans et al. 2006; Monteith et al.
2007), and changes in total solar UV radiation or an increase in temperature due to

65 Dissolved Organic Matter in Natural Waters
global warming (Freeman et al. 2001a; Zhang et al. 2010; Sinha et al. 2001; Sobek
et al. 2007; Rastogi et al. 2010). The increase in temperature driven by solar radia-
tion is effective in inducing photoinduced and microbial processes of OM (includ-ing DOM and POM) as well as in enhancing photosynthesis. This is consistent with data from the Central England Temperature Record (Parker et al. 1992), showing that mean summer temperatures across England were 0.66 °C higher during the 1990s than in the preceding 30 years. Model studies predict that the production of new DOM due to photosynthetic processes from winter to summer would vary from 6 to 60 %, due to a large seasonal variation in light intensity (Anderson and Williams 1998; Bratback and Thingstad 1985). The factors affecting the photosyn-
thesis in natural waters are discussed in detail in chapter “Photosynthesis in Nature: A New Look”.
6.7 Metal Ions Complexation and Salinity
Metal ions can complex the DOM functional groups (fulvic and humic acids of vascular plant origin, autochthonous fulvic acids of algal or phytoplankton origin, tryptophan, protein, algae and so on) and can induce structural changes (e.g. molec-ular conformation or rigidity) and formation of aggregates. Complexation would thus change the outer appearance of the molecule and its optical properties, such as absorption properties of CDOM and fluorescence properties of FDOM, either increasing or decreasing them (Mostofa et al. 2009a, 2011; Lead et al. 1999; Wang
and Guo 2000; Koukal et al. 2003; Mylon et al. 2003; Wu et al. 2004; Lamelas and Slaveykova 2007; Lamelas et al. 2009; Fletcher et al. 2010; Reiller and Brevet 2010; Sachs et al. 2010; Da Costa et al. 2011). Correspondingly, salinity can also affect the DOM components in seawater, both structurally and optically, modifying them in comparison to freshwater (Nakajima 2006; Blough et al. 1993; del Vecchio and Blough 2002; Boyd et al. 2010). Complexation of metal ions and the effect
of salinity are extensively discussed in Chapters “Photoinduced and Microbial Degradation of Dissolved Organic Matter in Natural Waters”, “Fluorescent Dissolved Organic Matter in Natural Waters”, and “Complexation of Dissolved Organic Matter with Trace Metal Ions in Natural Waters”.
6.8 Global Warming
Global warming may affect DOM in two ways: First, global warming could fur –
ther enhance atmospheric CO 2, because of elevated net primary productivity
and increases root exudation of DOC in soil environments (Freeman et al. 2001b, 2004; Lavoie et al. 2005; Fenner et al. 2007a, b; Wolf et al. 2007; Kang
et al. 2001; Tranvik and Jasson 2002; Monteith et al. 2007; Evans et al. 2002; Dorodnikov et al. 2011). This process ultimately leaches allochthonous DOM

66 K. M. G. Mostofa et al.
into the aquatic ecosystem. Second, global warming may accelerate the pho-
toinduced and microbial decomposition of DOM to produce compounds such as H
2O2, CO 2, DIC, NO 3−, PO 43−, NH 4+, LMWDOM and so on (Mostofa and
Sakugawa 2009; Johannessen and Miller 2001; Ma and Green 2004; Xie et al.
2004; Johannessen et al. 2007; Palenik and Morel 1988; Kotsyurbenko et al. 2001; Lovley 2006). The availability of these compounds can enhance photosyn-
thesis and ultimately increase the primary and secondary production. These pro-cesses can induce the formation of autochthonous DOM and are usually expected to deteriorate the quality of natural waters. All these processes are extensively discussed in chapter “Impacts of Global Warming on Biogeochemical Cycles in Natural Waters”.
7 Possible Mechanisms for Increased and Declined DOM
Contents in Surface Waters
Production of autochthonous DOM is a well-known phenomenon in stagnant
surface waters, particularly in lakes and oceans. The corresponding increase in autochthonous DOC is, on average, 0–102 % in lakes and 0–194 % in the oceans’ epilimnion compared to the hypolimnion during the summer stratifica-tion period (Table 2 ). Increased concentrations of DOC in surface waters are a
commonly observed phenomenon in North America and in North and Central Europe including UK, the Czech Republic, Finland, Norway, Canada, USA and so on (Hejzlar et al. 2003; Worrall et al. 2004a, 2007; Evans et al. 2005, 2006;
Monteith et al. 2007; Hongve et al. 2004; Monteith and Evans 2005; Larsen
et al. 2011; Clutterbuck and Yallop 2010; Yallop et al. 2010; Freeman et al. 2001b; Bouchard 1997; Skjelkvåle et al. 2001, 2005; Driscoll et al. 2003; Stoddard
et al. 2003; Vuorenmaa et al. 2006). The increase in DOC export also enhances the export of humic DOC from upland peat catchments (Yallop et al. 2010). It is estimated that the increase in mean DOC concentrations between the first and last 5 years of monitoring in UK’s Acid Waters Monitoring Network (AWMN) streams and lakes are 32–135 % in 11 streams and 31–140 % in 11 lakes (Evans
et al. 2006; Monteith and Evans 2005). DOC in UK rivers arises from a num-ber of sources including: decomposition of deep peat if present (McDonald et al. 1991), sewage (Eatherall et al. 2000), industrial point-source effluents (Tipping
et al. 1997) and products of early stages of plant decomposition (Palmer et al. 2001). The long-term trends of increasing or decreasing DOC concentration are not evident in various lakes except at the Experimental Lakes Area, where an increase in DOC is correlated with a decrease in summer total solar radiation and an increase in summer precipitation (Zhang et al. 2010). The initial DOM contents are also important to enhance the autochthonous production of DOC in aquatic ecosystems. Moreover, several mechanisms have been suggested to explain the enhancement of aquatic DOC including: increased terrestrial vegetation cover in response to climate change (Larsen et al. 2011; Worrall et al. 2003; Freeman

67 Dissolved Organic Matter in Natural Waters
et al. 2001b; Stoddard et al. 2003; Evans et al. 2005) and associated increases in
enchytraeid worm activity (Cole et al. 2002; Carrera et al. 2009); increasing CO 2-
mediated stimulation of primary productivity (Freeman et al. 2004); hydrological change (Hongve et al. 2004; Evans et al. 2005); artificial peat drainage (Worrall et al. 2003); the occurrence of severe drought events (Watts et al. 2001; Worrall and
Burt 2004); and the removal of decomposition-inhibiting phenolic compounds fol-
lowing prolonged water table drawdown (Freeman et al. 2001a). However, these mechanisms are not sufficiently well documented yet to understand the increased DOC concentrations in natural waters.
One of the possible mechanisms leading to increased DOC concentrations in
surface waters is the enhancement of photosynthesis. Two different processes are involved, depending on the sources of DOM.
(i) The first issue is that increased soil respiration may increase the decompo-
sition rates of soil OM due to the effect of global warming. Furthermore, elevated CO
2 enhances DOC supply in peat soils because of elevated net pri-
mary productivity and increased root exudation of DOC in soil environments (Freeman et al. 2001b, 2004; Lavoie et al. 2005; Fenner et al. 2007a, b; Wolf
et al. 2007; Kang et al. 2001; Tranvik and Jasson 2002; Monteith et al. 2007;
Evans et al. 2002; Dorodnikov et al. 2011). This process ultimately leaches
allochthonous DOM into the aquatic ecosystem. The increased activity of enchytraeid worms (the dominant invertebrates in upland peats) at higher temperature increases the microbial activity in peat and enhances nutrient mineralization (Cole et al. 2002). The mineralization of C- and N-containing compounds would increase the losses of nitrate and DOC (Cole et al. 2002).
(ii) The second issue is that the allochthonous DOM that is increasingly released into surface waters can undergo photoinduced decomposition to generate H
2O2, CO 2 and DIC (dissolved CO 2, H2CO3, HCO 3−, and CO 32-), or micro-
bial degradation with production of H 2O2, CO 2, DIC, CH 4, PO 43-, NH 4+
and so on (Lovley et al. 1996; Johannessen and Miller 2001; Ma and Green 2004; Xie et al. 2004; Johannessen et al. 2007; Palenik and Morel 1988; Clark
et al. 2004; Kotsyurbenko et al. 2001). Many of these compounds are able to enhance photosynthesis (Mostofa et al. 2009a; Komissarov 1994, 1995, 2003;
Li et al. 2011; Li 1994; Zubkov and Tarran 2008; Beardall et al. 2009a, b;
Wu and Gao 2009; Liu et al. 2010). This process can fuel primary and sec-ondary production, thereby leading to enhanced aquatic OM and DOM. In fact, algae and phytoplankton can produce autochthonous DOM under both photoinduced and microbial respiration or assimilation, which contributes to increasing DOM in natural waters. Under elevated carbon dioxide levels, the proportion of DOM derived from recently assimilated CO
2 is ten times
higher compared to the control cases (Freeman et al. 2004). In addition, new DOC release is far more sensitive to environmental drivers that affect net pri-mary productivity compared to decomposition alone (Freeman et al. 2004).

68 K. M. G. Mostofa et al.
Moreover, photosynthesis depends on several key factors that have already
been discussed and that could thus indirectly affect the occurrence of DOM in natural waters (see also chapter “Photosynthesis in Nature: A New Look”).
On the other hand, declined concentrations of DOC have been observed in sev-
eral surface waters including south west of England, northern Scandanavia and Italy (Worrall et al. 2004a, 2007; Schindler et al. 1996; Skjelkvåle et al. 2001;
Bertoni et al. 2010; Minella et al. 2011). Of the 315 catchments examined in the UK, 18 % (55 catchments) have shown significant decreases in DOC concentra-tion over the last 10 years (Worrall et al. 2007). DOC concentrations in the epilim-nion have decreased from 119 to 57 μM C (average values during 1980–1984 and 2000–2007, respectively). In lake Maggiore (Italy), chlorophyll a concentrations averaged 5.9 μg l
−1 in the period 1980–1990, decreased to 4.0 μg l−1 in the fol-
lowing decade (1990–2000) and underwent a further decrease (to 2.0 μg l−1) in
the period 2000–2007 (Bertoni et al. 2010). The observed DOC decline is pre-sumably caused by a decrease of total phosphorus and of the organic loadings to the lake, because of a decrease of the anthropic impact. The consequences are a decrease of in-lake productivity and pronounced changes in phytoplankton compo-sition, including higher biodiversity, reduced biovolume and lower average com-munity cell size (Bertoni et al. 1998, 2008; Callieri and Piscia 2002; Morabito and
OggioniA 2003; Salmaso et al. 2003; Rogora 2007).
The decline of DOC concentrations in surface waters would be linked to lower
photosynthesis and often to relatively low contents of DOM. The latter may also be the result of low precipitation, which generally decreases to input of soil alloch-thonous DOM to natural waters. Waters with low contents of DOM would produce low amounts of photoinduced or microbial end products, which may significantly decrease the primary and secondary production with a subsequent decline of autochthonous DOM (see also chapter “Photosynthesis in Nature: A New Look”). Low production of autochthonous DOM would further contribute to the decline of DOC in natural waters. In fact, production of autochthonous DOM may some-times offset the DOM decomposition by natural sunlight. For the same reason, soil inputs of allochthonous humic substances to surface waters during the summer stratification period may enhance photosynthesis and increase the autochthonous DOM in natural waters.
In some cases, drought can increase the DOM levels (Freeman et al. 2001a;
Worrall and Burt 2008; Vazquez et al. 2011; Evans et al. 1999; Worrall et al. 2006; Holden and Burt 2002, 2003). The mechanism behind this phenomenon is that
waters with high contents of DOM may undergo high photoinduced and microbial DOM degradation under drought conditions. The related production of photoin-duced and microbial end products may be responsible for enhancement of high photosynthesis and, therefore, of high primary and secondary production. The
latter phenomenon would ultimately lead to increased DOM contents.
The autochthonous production of DOM depends on several factors in natu-
ral waters, and particularly in lakes and oceans (Table 2) (Mostofa et al. 2005a, 2009a; Fu et al. 2010; Ogawa and Ogura 1992; Mitra et al. 2000; Yoshioka et al.

69 Dissolved Organic Matter in Natural Waters
2002a; Hayakawa et al. 2003, 2004; Anderson and Williams 1998; Bushaw et al.
1996; Takahashi et al. 1995; Marañòn et al. 2004). These factors can be summa-
rized as follows (Mostofa et al. 2009a): seasonal terrestrial riverine input; acidity or alkalinity, conductivity and pH, the variation of which indicates major differ –
ences in water chemistry that could influence photoreaction rates as well as the structure and speciation of organic matter. Other important factors are: anthropo-genic activities; water transparency and the related light penetration through the water column; standing stocks of carbon; stratification of the water column; stir –
ring/mixing of water by strong wind; photosynthetically active radiation (PAR) and water temperature; microbial degradation of DOM; ecosystem metabolism and bacterial growth; release of large amounts of dissolved organic compounds by wetland and littoral macrophytes; vertical mixing of the water column due to tem-perature effects; habitat structure and diversity.
8 Emerging Contaminants in Natural Waters
Emerging contaminants are generally detected in soil, sediment, air, water, aquatic biota including fish, wildlife, terrestrial earthworms, and humans (Richardson 2003, 2007; Mottaleb et al. 2005, 2009; Richardson and Ternes 2005, 2011; Buser et al.
2006; Schmid et al. 2007; Farré et al. 2008; Kinney et al. 2008; Guo et al. 2009; Ramirez et al. 2009; Citulski and Farahbakhsh 2010; Kumar and Xagoraraki 2010; Pal et al. 2010; Yoon et al. 2010; Kleywegt et al. 2011; Yu et al. 2011; Daughton and Ternes 1999; Keith et al. 2001; Heberer 2002; Balmer et al. 2004; Brooks et
al. 2005; Duedahl-Olesen et al. 2005). According to these studies, emerging con-
taminants (usually emerging organic contaminants) are typically defined as organic substances that occur in very small amount (usually at concentration levels of nano-grams to micrograms per liter), are persistent and have potential health effects on organisms including humans, fish and wildlife, or other adverse ecological effects. Such contaminants are either of anthropic origin (e.g. municipal, industrial, agricul-tural and human activities and waste water treatment processes) or naturally pro-duced, e.g. during the algal (or phytoplankton) blooms in surface water.
Emerging contaminants include a diverse group of organic compounds that
can be classified as pharmaceuticals, personal care products (PCPs), endocrine-disrupting compounds (EDCs), steroids and hormones, drinking water disinfection byproducts (DBPs), perfluorinated compounds (PFCs), brominated flame retardants (including polybrominated diphenyl ethers), sucralose and other artificial sweeten-ers, benzotriazoles, naphthenic acids, antimony, siloxanes, sunscreens/UV filters, musks, algal toxins, perchlorate, dioxane, pesticides, ionic liquids or organic salts, nanomaterials, gasoline additives and their transformation products, as well as microorganisms (Richardson 2003, 2007; Mottaleb et al. 2005, 2009; Richardson
and Ternes 2005, 2011; Buser et al. 2006; Schmid et al. 2007; Farré et al. 2008;
Kinney et al. 2008; Guo et al. 2009; Ramirez et al. 2009; Citulski and Farahbakhsh 2010; Kumar and Xagoraraki 2010; Pal et al. 2010; Yoon et al. 2010; Kleywegt

70 K. M. G. Mostofa et al.
et al. 2011; Yu et al. 2011; Daughton and Ternes 1999; Keith et al. 2001; Heberer
2002; Balmer et al. 2004; Brooks et al. 2005; Duedahl-Olesen et al. 2005).
It is estimated that approximately 3000 different substances are used as phar –
maceutical ingredients, including painkillers, antibiotics, antidiabetics, beta-blockers, contraceptives, lipid regulators, antidepressants and impotence drugs (Richardson and Ternes 2011). Pharmaceuticals that arise concern for possible
chronic toxicity are salicylic acid, diclofenac, propranolol, clofibric acid, carba-mazepine, atenolol, bezafibrate, cyclophosphamide, ciprofloxacin, furosemide, hydrochlorothiazide, ibuprofen, lincomycin, ofloxacin, ranitidine, salbutamol, sulfamethoxazole, diltiazem, acetaminophen, chloramphenicol, florfenicol, thia-mphenicol and fluoxetine (Pal et al. 2010; Richardson and Ternes 2011; Hoeger
et al. 2005; Carlsson et al. 2006; Pomati et al. 2006; Kim et al. 2007; Lai et al. 2009).
Occurences of various hormones in natural waters as priority drinking water contam-
inants are estriol [E3], estrone, progesterone, 17α -ethinylestradiol [EE2], 17α -estradiol,
17β-estradiol [E2], testosterone, androstenedione, equilenin, equilin, mestranol,
and norethindrone (Pal et al. 2010; Richardson and Ternes 2011). Synthetic musk compounds have diverse chemical structures, such as nitroaromatic groups includ-ing musk xylene (1-tert-butyl-3,5-dimethyl-2,4,6-trinitrobenzene) and musk ketone (4-tert-butyl-2,6-dimethyl-3,5-dinitroacetophenone); polycyclic structures including 7-acetyl-1,1,3,4,4,6-hexamethyl-1,2,3,4-tetrahydronaphthalene (AHTN; trade name, tonalide), 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-(g)-2-benzopyran (HHCB; trade name, galaxolide), 4-acetyl-6-tert-butyl-1,1-dimethylindan (ADBI; trade name, celestolide), dihydropentamethylindanone (DPMI; trade name, cashmeran), and 5-acetyl-1,1,2,3,3,6-hexamethylindan (AHMI, trade name phantolide). They are stud-ied in waters, waste sludge and air (Richardson and Ternes 2011; Gomez et al. 2009; Wombacher and Hornbuckle 2009; Clara et al. 2011; Ramirez et al. 2011). It has been shown that galaxolide is the most abundant musk detected in wastewater, reaching up to 2069 and 1432 ng L
−1 in influents and effluents, respectively.
Siloxanes include cyclic siloxanes, octamethylcyclotetrasiloxane (D4), deca-
methylcyclopentasiloxane (D5), dodecamethylcyclohexasiloxane (D6), tetra-decamethylcycloheptasiloxane (D7) and linear siloxanes (Richardson and Ternes 2011; Kierkegaard et al. 2011). Approximately 600 different pesticides are applied annually in the US, whilst in Japan more than 450 active products are dis-tributed annually among 5,400 commercial products. The two countries are key pesticide users in the world (Chen et al. 2007; Guo et al. 2009; Majewski et al.
2000; Derbalah et al. 2003; Qiu et al. 2005). Several pesticide degradation prod-ucts are also of concern, such as: alachlor ethanesulfonic acid (ESA), alachlor oxanilic acid (OA), acetochlor ESA, acetochlor OA, metolachlor ESA, meto-lachlor OA, 3-hydroxycarbofuran, terbufos sulfone, alachlor ESA and OA, ace-tochlor ESA and OA, metolachlor ESA and OA, thiophenol and phenyl disulfide from dyfonate hydrolysis; 4-chloro-2-methylphenol and 4-chloro-2-methyl-6-ni-trophenol from [(4-chloro-2-methylphenoxy)acetic acid] (MCPA) phototrans-formation; desphenyl-chloridazon (DPC) and methylated-DPC of N-chloridazon degradation (Richardson and Ternes 2011; Buttiglieri et al. 2009; Chiron et al. 2009; Wang et al. 2010).

71 Dissolved Organic Matter in Natural Waters
Benzotriazoles and other benzo-related contaminants are detected in water,
the most common being benzotriazole, tolyltriazole, benzothiazoles, and ben-
zosulfonamides in waters (Richardson and Ternes 2011; Jover et al. 2009; van Leerdam et al. 2009; Matamoros et al. 2010). Perfluorinated compounds (PFCs) include perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA) and a number of structurally-related compounds such as fluorinated telomer alco-hols (FTOHs), perfluorobutanoic acid (PFBA), perfluorohexanoic acid (PFHxS), Perfluorobutanesulfonate (PFBS), perfluoropropane sulfonate (PFPrS), perfluoro-ethane sulfonate (PFEtS), perfluorooctane sulfonamide (PFOSA), N-ethyl perfluo-rooctane sulfonamide acetate (N-EtFOSAA), perfluorododecanoic acid (PFDoDa), perfluoroundecanoic acid (PFUnDa), perfluorodecanoic acid (PFDA), perfluorono-nanoic acid (PFNA), perfluoroheptanoic acid (PFHpA), perfluorohexanoic acid (PFHxA), perfluoropentanoic acid (PFPeA), and perfluoropropanoic acid (PFPrA). They are mostly volatile and subject to metabolism or degradation that leads to the formation of their persistent sulfonate and carboxylic acid forms (Richardson and Ternes 2011; Andersen et al. 2008; Farré et al. 2008; Shi et al. 2008; Mak et al. 2009).
Perchlorate is mostly found as an impurity of sodium hypochlorite (liquid bleach)
and as ammonium perchlorate (Richardson and Ternes 2011). Ionic liquids are com-
posed of cationic or anionic polar headgroups with accompanying alkyl side chains. The cationic head groups include imidazolium, pyridinium, pyrrolidinium, morpho-linium, piperidium, quinolinium, quaternary ammonium, and quaternary phosphonium moieties. The anionic head groups include tetrafluoroborate (BF
4–), hexafluoro-
phosphate (PF 6–), bis(trifluoromethylsulfonyl)-imide [(CF 3SO2)2 N–], dicyanamide
[(CN) 2N–], chloride, and bromide (Richardson and Ternes 2011; Pham et al. 2010).
The chemical structures of nanomaterials are highly varied, including fuller –
enes, nanotubes, quantum dots, metal oxanes, TiO 2 nanoparticles, nanosilver, and
zerovalent iron nanoparticles (Richardson and Ternes 2011). Artificial sweeten-ers in natural waters mainly include sucralose, acesulfame, cyclamate, saccharin, aspartame, neotame, and neohesperidin dihydrochalcone (NHDEC) (Richardson and Ternes 2011; Scheurer et al. 2009).
Toxins of blue–green algal origin commonly occur as microcystins, nodularins,
anatoxins, cylindrospermopsin, and saxitoxins, while red tide toxins are detected as brevetoxins in natural waters (Richardson and Ternes 2011; dos Anjos et al. 2006; Wood et al. 2006; Zhao et al. 2006b; Smith et al. 2011). Saxitoxin variants recorded in cyanobacteria include decarbamoyl derivatives (dc), gonyautoxins (GTX); neosaxitoxin (neoSTX), N-sulphonocarbamoyl toxins (C-toxins), saxitoxin (STX) and a class of toxins produced by Lyngbya wollei (Humpage et al. 2010). Furthermore, chlorination of microcystin-LR and of cylindrospermopsin can give several byproducts, which are identified as chloro-microcystin, chloro-dihydroxy-microcystin, dichloro-dihydroxy-microcystin, trichlorohydroxy-microcystin, and several dihydroxy-microcystins (Merel et al. 2009).
Endocrine disrupting compounds or chemicals (EDCs) can disrupt the devel-
opment of the endocrine system and of organs that respond to endocrine signals in organisms. These organisms can be indirectly exposed during prenatal and/or early postnatal life and the effects of exposure during development are permanent

72 K. M. G. Mostofa et al.
and irreversible (Colborn 1993). In addition, transgenerational exposure can
result from the exposure of the mother to a chemical at any time throughout her life before producing offspring, due to persistence of EDCs in body fat. EDCs can then be mobilized during egg laying or pregnancy and lactation (Colborn 1993). EDCs include pesticides and their metabolites, DDT and its metabolites, pentachlorophenol (PCP), alkylphenols (e.g. penta- to nonylphenols, 4-tert-octylphenol) polychlorinated compounds (e.g. polychlorinated dibenzo-p-di-oxins, polychlorinated dibenzofurans, and polychlorinated biphenyls: PCBs), bisphenolic compounds (phenylphenol and bisphenol A), polybrominated diphe-nyl ethers, nematocides, naphthenic acids, pharmaceuticals (diethylstilbestrol), triclosan (2,4,4
′-trichloro-2 ′ 4′-trichloro-2 ′-hydroxydiphenyl ether), phthalate
esters (e.g. benzylbutyl phthalate, diethylhexyl phthalate, diphenylphthalate), ster –
oids (e.g. ethynyl estradiol, 17β-estradiol, diethylstilbestrol), natural estrogens (e.g. 17β-sitosterol, estriol, estrone), natural androgens (e.g. testosterone), natural phytoestrogens (soy, alfalfa and clover), naturally occurring compounds (lignans, coumestans, isofavones, mycotoxins), parabenes (hydroxybenzoate derivatives), organotins, and inorganic metal ions (cadmium, lead, mercury, antimony, uranium) (Citulski and Farahbakhsh 2010; Richardson and Ternes 2011; Colborn 1993; Bolger et al. 1998; Servos 1999; Petrović et al. 2001; Rhind 2002; Montgomery-Brown et al. 2003; Rudel et al. 2003; Ishibashi 2004; Kimura et al. 2004; Auriol et al. 2006; Diamanti-Kandarakis et al. 2009; Xu et al. 2011).
Microorganisms of concern are Cryptosporidium, Giardia, E. coli, pathogens, aer –
omonas, coliphages, viruses, total coliforms, Helicobacter pylori, enterococci, E. coli O157:H7 and H1N1 (swine flu) (Liu et al. 2009, 2010; Richardson and Ternes 2011;
Yañez et al. 2009; Li et al. 2010; Vikesland and Wigginton 2010; Wildeboer et al. 2010).
8.1 Sources of Emerging Contaminants in the Aquatic
Environment
Emerging contaminants are commonly derived from three major sources
(Richardson 2003, 2007; Mottaleb et al. 2005, 2009; Richardson and Ternes 2005,
2011; Buser et al. 2006; Schmid et al. 2007; Farré et al. 2008; Kinney et al. 2008; Guo et al. 2009; Ramirez et al. 2009; Citulski and Farahbakhsh 2010; Kumar and
Xagoraraki 2010; Pal et al. 2010; Yoon et al. 2010; Kleywegt et al. 2011; Yu et al.
2011; Daughton and Ternes 1999; Keith et al. 2001; Heberer 2002; Balmer et al.
2004; Brooks et al. 2005; Duedahl-Olesen et al. 2005): (i) anthropogenic sources
including atmospheric deposition but very often involving the effluents released by municipal, industrial, agricultural, and human activities, waste-water treatment processes, and so on; (ii) natural sources, including most notably the algal blooms in surface water; and (iii) photoinduced and microbial alteration of organic sub-stances during transport from rivers to lakes, oceans or other water sources.
The key point sources of pharmaceuticals are discharge of wastes and drugs
from hospitals; discharge of expired and consumed drugs from household;

73 Dissolved Organic Matter in Natural Waters
wastewater and solid wastes discharge from pharmaceutical industries; hormones
and antibiotics used in aquacultures; hormones and drugs used in livestock; com-pounds excreted from the human body in the form of non-metabolized parent molecules or as metabolites after ingestion and subsequent excretion, as well as the disposal of unused or expired medicinal products (Pal et al. 2010; Hernández et al. 2007; Pérez and Barceló 2007a, b; Nakada et al. 2008). The percentage of
parent compound from the human body is 6–39 or ≥70 % for various antibiotics, 6–39 or ≤5 % for analgesics and anti-inflammatory drugs, ≤5 % for antiepilep-
tic drugs (e.g. Carbamazepine), <0.5 or 50–90 % for beta-blockers, and 40–69 % for blood lipid regulators such as bezafibrate (Pal et al. 2010; Mompelat et al. 2009). The key entry route for pharmaceutical contaminants into natural waters is the point-source release from wastewater treatment plants (Daughton and Ternes 1999; Heberer 2002). Pharmaceuticals can also enter surface waters by run-off from fields treated with digested sludge (Farré et al. 2008). Veterinary drugs and their metabolites are transported through leaching or run-off from livestock slur –
ries when liquid manure is sprayed on agricultural field waters (Farré et al. 2008).
Personal care products (PCPs) (e.g. fragrances) can be discharged into aquatic
ecosystems through shower waste and finally from waste water treatment plants (Farré et al. 2008; Rimkus and Wolf 1996; Käfferlein et al. 1998; Smital et al. 2004; Peck 2006). UV filters used in sunscreens, cosmetics, and other PCPs are persistent in chlorinated water. Several halogenated by-products have been iden-tified, which can cause endocrine and developmental toxicity and estrogenicity (Kunz and Fent 2006; Negreira et al. 2008; Schmitt et al. 2008). Synthetic musk compounds are widely used as fragrance additives in many personal care prod-ucts, such as cleaning agents, air fresheners, house-hold products, perfumes, lotions, sunscreens, and laundry detergents (Richardson and Ternes 2011; Rimkus and Wolf 1996; Käfferlein et al. 1998; Smital et al. 2004). Steroids are excreted
in urine of humans as more hydrophilic glucuronides and sulfates, and free ster –
oids and conjugates are detected in sewage influent and effluent (Ascenzo et al. 2003; Reddy et al. 2005). Livestock wastes are potential sources of endocrine dis-rupting compounds and of steroidal estrogen hormones such as estradiol, estrone and estriol in natural waters (Raman et al. 2001; Hanselman et al. 2003; Furuichi
et al. 2006). Steroids were detected in more than 86 % of water samples from creeks where the cattle had direct access to the water (Kolodziej and Sedlak 2007).
Sources and pathways of xenohormone uptake by humans are mostly inhala-
tion (e.g., from indoor air), dermal absorption (e.g., from personal care products), and ingestion of food (Wagner and Oehlmann 2009). Another source of xenobiot-ics in foodstuff is the substances migrating from the packaging material, which can accumulate in the foodstuff. A variety of additives, such as stabilizers, antioxidants, coupling agents and pigments are used to optimize the properties of packaging mate-rials, which include for instance durability, elasticity and color (Lau and Wong 2000;
Casajuana and Lacorte 2003; Zygoura et al. 2005; Fankhauser-Noti et al. 2006).
Dioxane is a high production chemical that is used as solvent stabilizer in the
manufacture and processing of paper, cotton, textile products, automotive coolants, cosmetics and shampoos, as well as a stabilizer in 1,1,1-trichloroethane (TCA), a

74 K. M. G. Mostofa et al.
popular degreasing solvent (Richardson 2007; Richardson and Ternes 2011; Lee
et al. 2011). Siloxanes are widely used in PCPs and in a number of household
products, such as cosmetics, deodorants, soaps, hair conditioners, hair dyes, car waxes, baby pacifiers, cookware, cleaners, furniture polishes, and water-repellent windshield coatings (Richardson and Ternes 2011; Kierkegaard et al. 2011).
Ammonium perchlorate is used in solid propellants for rockets, missiles, fireworks
and highway flares. It can also be added in drinking water treatment processes as an impurity of sodium hypochlorite (liquid bleach), and is present as naturally-occuring perchlorate in fertilizers (e.g., Chilean nitrate). Formation of perchlorate can also take place upon reaction of chlorine radicals with ozone in the troposphere during the sum-mer periods (Richardson and Ternes 2011; Parker 2009; Furdui and Tomassini 2010).
Benzotriazoles are complexing agents that are mostly used as anticorrosives or
corrosion inhibitors (e.g., in engine coolants, aircraft deicers and antifreezing liq-uids), as UV-light stabilizers for plastics, for silver protection in dish-washing liquids, as anti-foggants in photography, and in aircraft de-icing/anti-icing fluids (ADAFs). These compounds are responsible for acute Microtox activity (Richardson 2007;
Lovley 2006; Cancilla et al. 1997). They are soluble in water, resistant to biodegrada-
tion, and are only partially removed in wastewater treatment (Richardson 2007).
Naphthenic acids are a complex mixture of alkylsubstituted acyclic and cyclo-
aliphatic carboxylic acids that dissolve in water at neutral or alkaline pH and have surfactant-like properties (Richardson and Ternes 2011). The key sources of naph-thenic acids are the residual tailing water left over from the extraction of crude oil from oil sands, coal deposits, and petroleum industries (Richardson 2007; Richardson and Ternes 2011; Headley et al. 2009; Scott et al. 2009). PFCs are
widely used in fabrics and carpets, paints, adhesives, waxes, polishes, metals, elec-tronics, fire-fighting foams and caulks, as well as grease-proof coatings for food packaging such as microwave popcorn bags, French fry boxes, hamburger wrap-pers, and so on (Richardson and Ternes 2011).
Pesticides are generally released from agricultural fields to rivers or nearby waters
by surface runoff, induced by either atmospheric precipitation or overflow and drain-age of agricultural field waters (Wang et al. 2007; Guo et al. 2009; Richards and Baker 1993; Majewski et al. 2000; Derbalah et al. 2003; Qiu et al. 2005). However, the new
sources of dichlorodiphenyltrichloroethane (DDT) are mostly connected with con-tinuing illegal applications of technical DDT, use of technical DDT-containing anti-fouling paint in commercial fishing boat maintenance, and presence of DDT residues in dicofol, although its use is internationally forbidden (Wang et al. 2007; Guo et al. 2009; Qiu et al. 2005). Nonylphenol polyethoxylates (NPEOs) and alkylphenol eth-oxylates (AEOs) are non-ionic surfactants, which are widely used in household, clean-ing products, paints, pesticides, and industrial processes such as paper and petroleum production (Farré et al. 2008; Fairchild et al. 1999; Strynar and Lindstrom 2008).
Ionic liquids have unique properties including tunable viscosity, miscibility,
and electrolytic conductivity. These properties make them useful for many appli-cations, such as organic synthesis and catalysis, production of fuel cells, batteries, coatings, oils, nanoparticles, as well as other chemical engineering and biotechnol-ogy applications (Richardson and Ternes 2011).

75 Dissolved Organic Matter in Natural Waters
Algal toxins are produced during algal or cyanobacterial blooms in natural
waters (Richardson and Ternes 2011; dos Anjos et al. 2006; Wood et al. 2006;
Zhao et al. 2006b; Smith et al. 2011).
Emerging pollutants can be altered in the environment by direct and indirect
photolysis, hydrolysis, other chemical processes, biodegradation, sorption, vola-tilization and dispersion, or by a combination of these processes. Environmental transformation can either contribute to the complete removal of the organic con-taminants, or produce transformation intermediates that can sometimes occur in the environment at higher levels than the parent compound and that can be as toxic or more toxic (Scully et al. 1988; Jensen and Helz 1998; Jameel and Helz 1999; Mitch et al. 2003; Strynar and Lindstrom 2008; Boxall et al. 2004; Gurr and Reinhard 2006; Jahan et al. 2008).
8.2 Transport of Emerging Contaminants in the Aquatic
Environment
Once released into the environment, emerging contaminants are transported into dif-
ferent aquatic organisms, sediments and plants, depending on the emission routes as well as their physico-chemical properties such as water solubility, vapor pressure and polarity (Guo et al. 2009; Richardson and Ternes 2011; Daughton and Ternes 1999;
Farré et al. 2008; Epel and Smital 2001). Emerging contaminants are generally per –
sistent, have a wide range of hydrophilicity/hydrophobicity, and many of them are liable to bioaccumulation and biomagnification in organisms and plants when present in the aquatic environment (Guo et al. 2009; Richardson and Ternes 2011; Daughton and Ternes 1999; Farré et al. 2008; Epel and Smital 2001). Aquatic organisms includ-
ing fish can accumulate emerging contaminants in certain body tissues. This phe-nomenon can take place either directly by bioaccumulation and biomagnification from water or by uptake of food such as OM (e.g. algae), sediments in water bed, small aquatic plants and so on, which have come in contact with the contaminants.
Emerging contaminants are mostly transmitted to humans through food con-
sumption, particularly fish and seafood (Wong et al. 2002; Meng et al. 2007). Synthetic musks are potential candidates as substrates or inhibitors of multixeno-biotic resistance (MXR) transporters (Daughton and Ternes 1999; Epel and Smital 2001). The multixenobiotic resistance (MXR) in aquatic organisms is mediated by the transport activity of transmembrane proteins belonging to the ATP-binding cassette (ABC) superfamily. These proteins are primarily involved in the active, ATP-dependent transport of biological molecules across plasma membranes (Smital et al. 2004; Higgins et al. 1988; Dean et al. 2001). The P-glycoprotein (P-gp) detected in ABC can transport drugs, xenobiotic compounds, antican-cer agents including the vinca alkaloids and anthracyclines, drugs against human immunodeficiency virus (HIV), fluorophores as well as typical environmental pol-lutants (Smital et al. 2004; Danø 1972; Juliano and Ling 1976; Smital and Kurelec 1998; Bard 2000; Litman et al. 2001). Various transmembrane transport proteins

76 K. M. G. Mostofa et al.
can thus cause a rapid efflux of a wide variety of potentially toxic xenobiotics out
of the cells of aquatic organisms. This is a ‘first line of defense’ against endog-enous and exogenous toxicants (Smital et al. 2004; Kurelec 1992; Epel 1998).
However, some environmental chemicals act as specific MXR inhibitors and
have the potential to block the active efflux of xenobiotics, thereby causing a sig-nificant increase of their intracellular accumulation. The main consequence of inhibition is an increase in chemosensitivity of aquatic organisms toward the many xenobiotics that are typically present in aquatic environments (Smital et al. 2004). Based on these considerations, otherwise innocuous environmental chemicals can be seen as a new class of environmentally hazardous chemicals that are termed as MXR inhibitors or chemosensitizers (Smital et al. 2004).
Perchlorate is a very water-soluble and environmentally stable anion, which can
accumulate in plants (including lettuce, wheat, and alfalfa) and can thus contribute to exposure in humans and animals (Richardson and Ternes 2011).
8.2.1 Toxicological Impacts of Emerging Contaminants
Emerging contaminants and their transformation byproducts have adverse effects
on the health of aquatic organisms (including algae, bacteria and fish), of animals and humans, as well as aquatic ecological effects (Guo et al. 2009; Kumar and Xagoraraki 2010; Pal et al. 2010; Derbalah et al. 2003; Scully et al. 1988; Jensen
and Helz 1998; Jameel and Helz 1999; Mitch et al. 2003; Pomati et al. 2006; Farré et al. 2008; Fairchild et al. 1999; Boxall et al. 2004; Jahan et al. 2008; McLeese et al. 1981; Ahel et al. 1987; Tyler et al. 1998; Scott and Jones 2000; Oberdorster and Cheek 2001; Cleuvers 2004; Ferrari et al. 2004; Bedner and MacCrehan 2006; Owen et al. 2007). Their effects can be characterized using seven attributes: preva-lence, frequency of detection, removal, bioaccumulation, ecotoxicity (for fish, daph-nid, and algae aquatic indicator species), pregnancy effects, and health effects. The latter attribute was characterized using seven sub-attributes: carcinogenicity, muta-genicity, impairment of fertility, central nervous system action, endocrine effects, immunotoxicity, and developmental effects (Kumar and Xagoraraki 2010).
Production of byproducts such as trihalomethanes (THMs), N-nitrosodimeth-
ylamine (NDMA), and organic chloramines in conventional and advanced wastewater treatment plants arises considerable concern. These compounds are in fact extremely toxic and carcinogenic to human beings and aquatic organ-isms, and have been found in drinking and natural waters (Scully et al. 1988; Jensen and Helz 1998; Jameel and Helz 1999; Mitch et al. 2003; Farré et al. 2008). Transformation products of some organics are often more persistent than the corresponding parent compounds, and can cause greater toxicity (Boxall
et al. 2004). For example, the major biodegradation product of nonylphenol ethox-ylates, nonylphenol, is much more persistent than the parent compounds and has estrogenic properties (Jahan et al. 2008). Pharmaceuticals and their transformation byproducts show acute toxicity to bacteria, algae, invertebrates, fish, mussels, and human embryonic cells (Guo et al. 2009; Kumar and Xagoraraki 2010; Pal et al.

77 Dissolved Organic Matter in Natural Waters
2010; Derbalah et al. 2003; Scully et al. 1988; Jensen and Helz 1998; Jameel and
Helz 1999; Mitch et al. 2003; Pomati et al. 2006; Farré et al. 2008; Fairchild et al.
1999; Boxall et al. 2004; Jahan et al. 2008; McLeese et al. 1981; Ahel et al. 1987; Tyler et al. 1998; Scott and Jones 2000; Oberdorster and Cheek 2001; Cleuvers 2004; Ferrari et al. 2004; Bedner and MacCrehan 2006; Owen et al. 2007). It has been shown that low part per trillion (10–100 ng L
−1) concentrations of steroidal
estrogen hormones can adversely affect the reproductive biology of aquatic wild-life such as fish, turtles and frogs, by disrupting the normal function of their endo-crine systems (Tyler et al. 1998; Oberdorster and Cheek 2001). The sex hormones (mainly estrogens and androgens) are of very high potential concern, followed by cardiovascular drugs, antibiotics and antineoplastics, the latter being used to cure abnormal tissue growth (neoplasms) (Sanderson et al. 2004).
Ethylene dibromide (EDB) is among the most commonly detected contami-
nants in groundwater. It is classified as a probable human carcinogen and is highly persistent in water (Richardson 2007). 1,4-Dioxane is also a widespread contami-nant in groundwater and is a probable human carcinogen (Richardson 2007). The transformation intermediates of nonylphenol ethoxylates (NPEOs) and alcohol ethoxylates (AEOs), in addition to the endocrine disrupting properties, are highly toxic and refractory and can cause hazards to aquatic ecosystems (Derbalah et al. 2003; Fairchild et al. 1999; McLeese et al. 1981; Ahel et al. 1987; Scott and Jones
2000). DDT and its metabolites can damage the nervous system, reproductive sys-tem, and liver. It is also a potential human carcinogen that can cause liver cancer (Guo et al. 2009). Ionic liquids are toxic, and their toxicity can widely vary among organisms and trophic levels (Pham et al. 2010).
EDCs have potential effects on organisms (microorganisms, wildlife, animals,
and humans) including: androgenic, estrogenic, anti-estrogenic and anti-andro-genic properties; disruption of the development of vital systems such as the endro-crine, reproductive, immune, and thyroid functions; sexual differentiation of the brain during fetal development; cognitive and motor function. Many of them are also suspected carcinogens (Richardson and Ternes 2011; Colborn 1993; Rhind 2002; Jansen et al. 1993; Nimrod and Benson 1996; Hansen 1998; Langer et al. 1998; Helleday et al. 1999; Vine et al. 2000; Moore et al. 2001; Fenton 2006). EDCs can have transient and persistent effects on mammary gland development depending on dose, exposure parameters and on whether exposure occurs during critical periods of gland growth or differentiation (Fenton 2006). Adverse effects from these abnormal developmental patterns include the presence of carcin-ogen-sensitive structures in the gland, in greater numbers or for longer periods. Inhibited functional differentiation can also be observed, leading to malnutrition or increased mortality of the offspring (Fenton 2006). Individually, adverse effects of EDCs exposure are detected on sperm production in rats and humans and reduc-tions in Sertoli cell number in sheep (Carlsen et al. 1995; Toppari et al. 1996; Lee et al. 1999; Sweeney et al. 2000). Reductions in embryo survival and consequent effects on the reproductive rate in females are observed for many mammalian and bird species (IEH) (IEH 1999). Finally, human health is adversely affected by con-sumption of food contaminated by EDCs.

78 K. M. G. Mostofa et al.
Currently, algal toxins or red tide toxins produced during algal blooms in lakes,
estuaries and oceans are responsible for adverse effects, including the increasing
incidence of loss of phytoplankton competitor motility, inhibition of photosynthe-sis and of enzymes, membrane damage, large fish kills, shellfish poisoning, deaths of livestock and wildlife, as well as illness and death in humans associated with the consumption of contaminated shellfish (Richardson 2007; Prince et al. 2008; Negri et al. 1995; Landsberg 2002; Legrand et al. 2003; Llewellyn 2006; Etheridge 2010). It has been shown that saxitoxin and its analogues are the only neurotoxins identified in Anabaena circinalis from the Murray Darling River. There an exten-sive A. circinalis bloom in 1991 resulted in the death of over 1600 stock (Humpage
et al. 1994; Bowling and Baker 1996; Steffensen et al. 1999). The mechanisms behind the effects of harmful algal blooms on organisms will be discussed in chap-ter “Photosynthesis in Nature: A New Look”. Finally, microorganisms are respon-sible for outbreaks of waterborne illness that have killed millions of people over the last few decades all over the world (Richardson and Ternes 2011).
8.3 Methodologies and Techniques of Emerging
Contaminants (Pharmaceuticals and Personal
Care Products) Detection
The liquid chromatography-tandem mass spectrometry (LC–MS/MS) screen-
ing method described here has been developed to target 23 pharmaceuticals and 2 metabolites with differing physicochemical properties in fish tissue by Ramirez and his colleagues (Ramirez et al. 2007). In this method, analysis of pharmaceuti-cals and their metabolites in fish tissue is conducted using reversed-phase separa-tion of target compounds with a C18 column and a nonlinear gradient, consisting of 0.1 % (v/v) formic acid and methanol. A 1:1 mixture of 0.1 M aqueous acetic acid (pH 4) and methanol is identified as optimal, resulting in extraction recoveries for 24 of 25 compounds exceeding 60 % among 10 solvents tested. Eluted ana-lytes are then introduced into the mass analyzer using positive or negative electro-spray ionization. Note that moderate-polarity solvents are generally observed to be most effective at removing target analytes from tissue.
Sample collection and preservation (Ramirez et al. 2007): Fish (Lepomis sp.)
were sampled from Pecan Creek (impacted by pharmaceuticals) and Clear Creek (not impacted by contaminants) streams to serve as test and reference specimens, respectively. Lateral fillets were dissected from fish collected at both sites and homogenized using a Tissuemiser (Fisher Scientific, Fair Lawn, NJ) set to rotate at 30,000 rpm. Pecan creek homogenates were stored individually, while Clear Creek homogenates were composited into a single sample. All tissues were stored at −20 °C prior to analysis.
Preparation of analytical sample (Ramirez et al. 2007): Approximately
1.0 g of tissue was combined with 8 mL of extraction solvent [a 1:1 mixture of

79 Dissolved Organic Matter in Natural Waters
0.1 M aqueous acetic acid (pH 4) and methanol] in a 20mL borosilicate glass vial
(Wheaton; VWR Scientific, Rockwood, TN), and the mixture was homogenized using a Tissuemiser (Fisher Scientific) at 30,000 rpm. Five surrogates (100.0 μ g/
mL in acetonitrile) were added to each sample: acetaminophen-d
4 (454 ng), fluox-
etine-d 6 (636 ng), diphenhydramine-d 3 (8.9 ng), carbamazepine-d 10 (38.5 ng),
and ibuprofen-13C3 (789 ng). Samples were shaken vigorously and mixed on a
rotary extractor for 5 min. Following extraction, samples were rinsed into 50-mL polypropylene copolymer round-bottomed centrifuge tubes (Nalge Co.; Nalgene Brand Products, Rochester, New York) using 1 mL of extraction solvent and cen-trifuged at 16,000 rpm for 40 min at 4 °C. The supernatant was decanted into 18-mL disposable borosilicate glass culture tubes (VWR Scientific), and the solvent was evaporated to dryness under a stream of nitrogen at 45 °C using a Zymark Turbovap LC concentration workstation (Zymark Corp., Hopkinton, MA). Samples were reconstituted in 1 mL of mobile phase, and a constant amount of the internal standards 7-aminoflunitrazapam-d
7 (100 ng) and meclofenamic
acid (1000 ng) was added. Prior to analysis, samples were sonicated for 1 min and filtered using Pall Acrodisc hydrophobic Teflon Supor membrane syringe filters (13 mm diameter; 0.2 μ m pore size; VWR Scientific, Suwanee, GA).
LC–MS/MS detection (Ramirez et al. 2007): A Varian ProStar model 210
binary pump equipped with a model 410 autosampler was used to detect the ana-lytes, which were separated on a 15 cm × 2.1 mm (5 μm, 80 Å) Extend-C18 col-umn (Agilent Technologies, Palo Alto, CA) connected with an Extend-C18 guard cartridge 12.5 mm × 2.1 mm (5 μm, 80 Å) (Agilent Technologies). A binary gradient consisting of 0.1 % (v/v) formic acid in water and 100 % methanol was employed to achieve chromatographic separation, whereas the time-scheduled elution program was as follows (min): 0, 2, 7, 12, 21 28, 34, 45, 50, 51, 65. The mobile-phase composition for 0.1 % formic acid was 93, 93, 85, 85, 52, 52, 41, 2, 2, 93, 93 and for methanol was 7, 7, 15, 15, 48, 48, 59, 98, 98, 7, 7, respectively. Additional chromatographic parameters were as follows: injection volume, 10 μL; column temperature, 30 °C; flow rate, 350 μL min
−1. Eluted analytes were moni-
tored by MS/MS using a Varian model 1200L triple-quadrupole mass analyzer equipped with an electrospray interface (ESI).
Each compound was infused individually into the mass spectrometer at a con-
centration of 1 μg mL
−1 in aqueous 0.1 % (v/v) formic acid at a flow rate of 10
μL min−1 for determining the best ionization mode (ESI + or −) and optimal
MS/MS transitions for target analytes. All analytes were initially tested using both positive and negative ionization modes while the first quadrupole was scanned from m/z 50 to [M + 100]. This can enable identification of the optimal source
polarity and the most intense precursor ion for each compound. Once these param-eters have been defined, the energy at the collision cell was varied, while the third quadrupole was scanned to identify and optimize the intensity of product ions for each compound. Additional instrumental parameters held constant for all analytes were as follows: nebulizing gas, N
2 at 60 psi; drying gas, N 2 at 19 psi; tempera-
ture, 300 °C; needle voltage, 5000 V ESI+ , 4500 V ESI-; declustering potential,
40 V; collision gas, argon at 2.0 mTorr.

80 K. M. G. Mostofa et al.
Extraction Recoveries (Ramirez et al. 2007): All samples were analyzed by
LC–MS/MS, and individual analyte recoveries were calculated using the following
equation:
where AX1, AIS1, AX2, and A IS2 represent peak areas for the analyte (X) and inter –
nal standard (IS) in groups 1 and 2 samples, respectively.
Identification of Pharmaceuticals Using LC–MS/MS
(Ramirez et al. 2007):
A LC–MS/MS total ion chromatogram resulting from analysis of clean
tissue (non-affected by contaminants) spiked with a mixture of stand-ard pharmaceuticals is depicted in Fig. 3. Peak identifications for pharma-
ceuticals in the chromatogram are as follows: (1) acetaminophen-d
4, (2)
acetaminophen, (3) atenolol, (4) cimetidine, (5) codeine, (6) 1,7-dimeth-ylxanthine, (7) lincomycin, (8) trimethoprim, (9) thiabendazole, (10) caf-feine, (11) sulfamethoxazole, (12) 7-aminoflunitrazepam-d
7 (+IS), (13)
metoprolol, (14) propranolol, (15) diphenhydramine-d 3, (16) diphenhydramine,
(17) diltiazem, (18) carbamazepine-d 10, (19) carbamazepine, (20) tylosin, (21)
fluoxetine-d 6, (22) fluoxetine, (23) norfluoxetine, (24) sertraline, (25) erythromy-
cin, (26) clofibric acid, (27) warfarin, (28) miconazole, (29) ibuprofen-13C3, (30)
ibuprofen, (31) meclofenamic acid (-IS), and (32) gemfibrozil. Three factors were presumably considered in selecting the target analytes (Table 3): First, number of prescriptions dispensed in the United States during 2005 (RxList 2005). Second, variability in structure, physicochemical properties, and therapeutic use. Third, rel-ative frequency of occurrence in soils, sediments, and biosolids. The frequency of detection of various PPCPs in analyzed sediment, soil, and biosolid samples (64–100 %) is typically much higher than in water (5 %). This may be due to variation in physicochemical properties favoring compound partitioning from water to solid environmental matrixes. Compounds residing in sediment may then be taken up by aquatic organisms via ingestion (Furlong et al. 2004; Brooks et al. 2005; Ramirez et al. 2007).
Optimized MS/MS transitions and collision energies employed for detection and
quantitation of each analyte are presented in Table 3 , along with the molecular struc-
ture and most common therapeutic use for each analyte. With the exception of eryth-romycin, selected precursors represent the molecular ion [M + H]
+ or [M − H]− for
each analyte. The most abundant precursor for erythromycin was found to be the [M + H − H
2O]+ ion at m /z 716. Selected product ions generally represent the most
abundant fragment observed for each precursor at the noted collision energy. Once suitable MS/MS transitions have been identified for each analyte, an aqueous mix-ture of reference standards was employed to optimize chromatographic parameters. A nonlinear gradient consisting of 0.1 % (v/v) formic acid and methanol resulted in near baseline resolution of the majority of analytes in ~50 min (Fig. 3 ). A 15-min Recovery =(Ax1/AIS1)/(Ax2/AIS2)×100%

81 Dissolved Organic Matter in Natural Waters
isocratic hold (93:7 formic acid–methanol) was added to the end of each run to
allow for column equilibration between injections. While the majority of analytes
were eluted as single peaks, erythromycin was consistently eluted as two partially
resolved peaks, which are attributed to differing retention characteristics for pre –
sumed Stereoisomers (Vanderford et al. 2003; Yang and Carlson 2004). In addition,
isotope effects on retention behavior were often observed for carbamazepine- d10 and
fluoxetine- d6 (peaks 18 and 19, Fig. 3). The observed retention time for carbamaze –
pine- d10 (30.08 min) was shorter than that observed for carbamazepine (30.53 min)
by almost 30 s. Correspondingly, a 20-s difference in retention time was observed
for fluoxetine- d6 (34.58 min) relative to that observed for fluoxetine (34.93 min),
although it is not evident in Fig. 3 due to coelution of norfluoxetine (35.13 min).
Finally, isotope effects were not observed for acetaminophen (peaks 1 and 2) and
diphenhydramine (peaks 15 and 16, Fig. 3) due to a lower degree of deuterium sub –
stitution and decreased resolution at shorter retention times. Four compounds such as
diphenhydramine, diltiazem, carbamazepine, and norfluoxetine were detected in fish
environmental samples (affected by contaminants; Fig. 4a), which were confirmed by
comparing the results of the fish samples unaffected by contaminants and spiked with
known amounts of their respective standards (Fig. 4b). The concentrations of these
pharmaceuticals were 0.66–1.32 ng g−1 for diphenhydramine, 0.11–0.27 ng g−1 for
diltiazem, 0.83–1.44 ng g−1 for carbamazepine, and 3.49–5.14 ng g−1 for norfluox –
etine detected in 11 of 11 contaminated environmental fish samples.
Fig. 3 LC-MS/MS total ion chromatogram resulting from analysis of clean tissue spiked with
a mixture of pharmaceutical standards. Peak identifications are as follows: (1) acetaminophen-
d4, (2) acetaminophen, (3) atenolol, (4) cimetidine, (5) codeine, (6) 1,7-dimethylxanthine, (7)
lincomycin, (8) trimethoprim, (9) thiabendazole, (10) caffeine, (11) sulfamethoxazole, (12) 7
aminoflunitrazepam-d 7 (+IS), (13) metoprolol, (14) propranolol, (15) diphenhydramine-d 3,
(16) diphenhydramine, (17) diltiazem, (18) carbamazepine-d 10, (19) carbamazepine, (20) tylo –
sin, (21) fluoxetine-d 6, (22) fluoxetine, (23) norfluoxetine, (24) sertraline, (25) erythromycin,
(26) clofibric acid, (27) warfarin, (28) miconazole, (29) ibuprofen-13C3, (30) ibuprofen, (31)
meclofenamic acid (-IS), and (32) gemfibrozil. Data source Ramirez et al. ( 2007 )

82 K. M. G. Mostofa et al.Table 3 Analyte-dependent mass spectrometry parameters for target compounds (data source Ramirez et al. 2007 ; RxList (The Internet Drug Index) 2005 )
Compound Use Structure Precursor ion Collision energy (eV) Product ion pKaa
ESI positive analytes
Acetaminophen Analgesic 152 [M + H]+−11.0 110 9.86
Atenolol Anti-hypertension 267 [M + H|+−21.5 145 9.16
Cimetidine Anti-acid reflux 253 [M + H]+−13.5 159 7.07
Codeine Analgesic 300 [M + Η]+−38.0 215 8.25
1,7-dimethylxanthine Caffeine
metabolite181 [M + H]+−15.5 124 8.50
Liltcomycin Antibiotic 407 [Μ + H]+−15.5 359 8.78
(continued)

83 Dissolved Organic Matter in Natural Waters
Compound Use Structure Precursor ion Collision energy (eV) Product ion pKaa
Trimethoprim Antibiotic 291 [Μ + Η]+−17.5 261 7.20
Thiabendazole Antibiotic 202 [M + H]+−23.0 175
Caffeine Stimulant 195 [Μ + Η]+−16.0 138
Sulfamethoxazole Antibiotic 254 [Μ + Η]+−13.0 156 5.81
Metoprelol Anti-hypertension 268 [M + H]+−15.5 191 9.17
Propranolol Anti-hypertension 260 [Μ + H]+−11.0 116 9.14
Diphenhydramine Antihistamine 256 [Μ + H]+−11.5 167 8.76
Diltiazem Anti-hypertension 415 [M + H]+−22.0 178 8.94
Table 3 (continued)
(continued)

84 K. M. G. Mostofa et al.Compound Use Structure Precursor ion Collision energy (eV) Product ion pKaa
Carbamazepine Anti-seizure 237 [Μ + Η]+−13.5 194
Tylosin Antibiotic 916 [M + H]+−31.5 174 7.39
Fluoxetine Anti depressant 310 [M + H]+−6.0 148 10.1
Norfluoxetine Fluoxetine metabo-
lite296 [M + H]+−45 134 9.05
Sertraline Antidepressant 306 [M + H]+−11.0 275 9.47
Erythromycin Antibiotic 716 [M + H − H2O]+−18.0 558 8.16
Table 3 (continued)
(continued)

85 Dissolved Organic Matter in Natural Waters
Compound Use Structure Precursor ion Collision energy (eV) Product ion pKaa
Warfarin Anli-coaulant 309 [M + H]+−14.0 163 4.50
Miconazole Antibiotic 417 [M + H]+−27.5 161 6.67
ESI Negative Analytes
Clofibric Acid Antilipemi 213 [M − H]−15.4 127 3.18
Ibuprofen Analgesic 205 [M − H|−7.0 161 4.41
Gemfibrozil Antilipemie 249 [M − H]−13.0 121 4.75
aCalculated values obtained from SciFinder database (@ 2006 American Chemical Society)
Table 3 (continued)

86 K. M. G. Mostofa et al.
Identification of Personal Care Products Using GC–MS/MS (Mottaleb
et al. 2009 )
Two gas chromatography-mass spectrometry (GC–MS) methods have been
described for simultaneous determination in fish of ten extensively used per –
sonal care products (PCPs) such as benzophenone, 4-methylbenxylidine camphor
Fig. 4 LC-MS/MS reconstituted ion chromatograms displaying analyte-specific quantitation and
qualifier ions monitored for ( a) a tissue extract from a fish (Lepomis sp.) that is affected by con –
taminants and ( b) an extract from fish tissue (not affected by contaminants) spiked with known
amounts of diphenhydramine (1.6 ng g−1), diltiazem (2.4 ng g−1), carbamazepine (16 ng g−1),
and norfluoxetine (80 ng g−1). The higher m/z fragment is more intense in all cases. Data source
Ramirez et al. ( 2007 )

87 Dissolved Organic Matter in Natural Waters
(4-MBC), m-toluamide, galaxolide, tonalide, musk xylene, musk ketone, celes-
tolide, triclosan, octocrylene and two alkylphenol surfactants such as p-octyl-phenol and p-nonylphenol. These methods consisted of extraction, clean-up, derivatization and analysis by gas chromatography–mass spectrometry with selected ion monitoring (GC–SIM–MS) or gas chromatography–tandem mass spectrometry (GC–MS/MS) techniques (Mottaleb et al. 2009). To assess recovery of target compounds from 1-g tissue homogenates, acetone was selected as opti-mal solvent for extracting compounds with dissimilar physicochemical properties from fish tissue. Initial experiments confirmed that GC–SIM–MS could be applied for analysis of lean fillet tissue (<1 % lipid) without gel-permeation chromatog-raphy (GPC), and this approach was applied to assess the presence of target ana-lytes in fish fillets collected from a regional effluent-dominated stream in Texas, USA. Benzophenone, galaxolide, tonalide, and triclosan were detected in 11 of 11 environmental samples at concentrations ranging from 37 to 90, 234 to 970, 26 to 97, and 17 to 31 ng g
−1, respectively. However, performance of this analytical
approach declined appreciably with increasing lipid content of analyzed tissues. Successful analysis of samples with increased lipid content was enabled by add-ing GPC to the sample preparation protocol and monitoring analytes with tandem mass spectrometry. Both analytical approaches were validated using fortified fil-let tissue collected from locations expected to be minimally impacted by anthro-pogenic influences. Average analyte recoveries ranged from 87 % to 114 % with
RSDs <11 % and from 54 % to 107 % with RSDs <20 % for fish tissue contain-ing <1 % and 4.9 % lipid, respectively. Statistically derived method detection lim-its (MDLs) for GC–SIM–MS and GC–MS/MS methodologies ranged from 2.4 to 16 ng g
−1, and from 5.1 to 397 ng g−1, respectively (Mottaleb et al. 2009). In
a following study, improvement of the MDL has been observed between 12 and 38 ng g
−1 by the GC–MS/MS methodology for the same PCPs using 2.0–2.5 g of
fish (Subedi et al. 2011).
9 Does DOM Act as Energy Source for Living Organisms
and Aquatic Ecosystem?
The concentration levels of DOC in groundwater are very variable: they reach
16–424 μM C in Asia, 42–15333 μ M C in Europe, 8–2333 μ M C in North America,
1108 ± 217–14167 ± 6333 μ M C in Botswana (Africa), 100–3000 μ M C in Brazil
(South America) (Table 2) (Mostofa et al. 2007a, Mostofa KMG et al., unpub-
lished data; Buckau et al. 2000; Bertilsson et al. 1999; McIntyre et al. 2005; Meier et al. 2004; Crandall et al. 1999; Schwede-Thomas et al. 2005; Pabich et al. 2001;
Michalzik et al. 2001; Anawar et al. 2002; Richey et al. 2002; Bradley et al. 2007). Groundwater is the main source of drinking water for many developing and devel-oped countries, including the USA. Groundwater has the advantage over surface water of being usually free of suspended solids, bacteria and other disease-causing microorganisms (Mostofa et al. 2009a). Interestingly, upland areas make up 30 %

88 K. M. G. Mostofa et al.
of the surface of Great Britain, but supply over 70 % of its drinking water (Watts
et al. 2001). Therefore, all the people uptake a certain amount of DOC everyday from
drinking water. According to the level of DOC in groundwater and considering an average water intake of 2 liters per day for adults (which can rise to ~5 liters for man-ual labor at high temperature), on average, every person intakes per day ~50–800 μ M
C in Asia, ~100–30000 μ M C in Europe and 20–5000 μ M C in the U.S.A. The inter –
esting question that arises is that these DOC contents are significant energy sources for human beings and for the other living organisms. Before addressing this question, it is important to examine which substances make up DOM in natural waters.
The contribution of humic substances (hydrophobic acids) in groundwater is
very variable in different countries, and is approximately included in the range of 12–98 % (1–80 % of fulvic acid and 2–97 % of humic acid). The contribu-tion of hydrophilic fractions is 1–82 % (Buckau et al. 2000; Bertilsson et al. 1999; Peuravuori and Pihlaja 1999; Leenheer et al. 1974; Thurman 1985c; Ford and
Naiman 1989; Schiff et al. 1990; Wassenaar et al. 1990; Malcolm 1991; Grǿn et al. 1996; Christensen et al. 1998; McIntyre et al. 2005; Mladenov et al. 2008). Along
with the humic substances, hydrophilic compounds (acidic, basic and neutral) and carbohydrates (mainly polysaccharides, ~1–10 %) are also present in groundwater (Thurman 1985a; Peuravuori and Pihlaja 1999; Artinger et al. 2000). The intake
of DOC by every person is approximately 20–30000 μM C, or 0.2–360 mg C L
−1
per day, for the average hydration of a human body in the case of groundwater.
It is generally well-known that carbohydrates can produce energy for all living
organisms. The sources of carbohydrates and humic substances are the same vas-cular plant material. DOM with its content of organic C and N is a thermodynamic anomaly that provides a major source of energy to drive aquatic and terrestrial ecosystems (Tranvik 1992; Salonen et al. 1992; Wetzel 1984, 1992; Hedges et al.
2000; Berner 1989). Energy changes (±) such as supply (+) or consumption (−)
of energy in the aquatic environment generally occur during the photoinduced and microbial degradation of DOM and organic matter, during the microbial loop and the photosynthesis (Mostofa et al. 2009a; Komissarov 1994, 1995, 2003; Miller
and Moran 1997; Li et al. 2011; Sherr and Sherr 1989; Carrick et al. 1991; Jones
1992; Tranvik 1992; Wetzel 1984, 1992). In addition, terrestrial DOM represents
a source of allochthonous energy for heterotrophs in receiving lakes, rivers, reser –
voirs, estuaries and coastal oceans (Mostofa et al. 2009a; Wetzel 1992; Smith and
Hollibaugh 1993; Kemp et al. 1997; Pace et al. 2004; Aller and Blair 2006). It has
been shown that DOM makes up 47 % of the energy which enters and 70 % of the energy which leaves the groundwater ecosystem (Fisher and Likens 1973). It has also been shown that undisturbed groundwater basins export only small amounts of energy (~1 %) from the upland regions, while the ramaining 99 % of forest production is consumed terrestrially (Fisher and Likens 1972, 1973). It is therefore
concluded that the DOM including humic substances can act as energy source and are vital for all living organisms (Mostofa et al. 2009a). Note that DOM in drink-
ing water can play a negligible energetic role for humans, due to the uptake of a substantially lower amount of organic carbon compared to foods (e.g. boiled rice, vegetables, fish, meat and so on) and beverages (e.g. fruit juices, alcohol, etc.).

89 Dissolved Organic Matter in Natural Waters
Humic substances (humic and fuvic acids) are extensively applied as bio-
medicines to decrease the gastric damage induced by ethanol, to protect organ-
isms against cell-wall disruption, to maintain antibacterial and antiviral properties, decrease viral respiratory illness, and to protect against cancer and related can-cer-causing viruses (Brzozowski et al. 1994; Klöcking et al. 2002; Peña-Méndez
et al. 2005). On the other hand, humic acid is a toxic factor for many mammalian cells and can be involved in the so-called humic acid-induced cytotoxicity (Peña-Méndez et al. 2005; Ho et al. 2003).
10 Scope of the Future Research
After the development of an effective method for TOC analysis in 1988, DOM has been mostly determined in developed countries since 1990 to date, but fewer studies have been carried out in developing countries. Considering the impor –
tance of DOM, it is important to determine its levels in natural water in develop-ing countries, also considering that the DOC concentrations in many watersheds have changed (either increased or decreased) over the last few decades. Moreover, emerging contaminants and their transformation byproducts are extensively exam-ined currently, but only limited information is available on their ecotoxicological impacts on the aquatic environments.
Some important research demands for future challenges are the follow-
ing: (i) Determination of concentration levels of DOM in important rivers and lakes in developing countries. (ii) Extraction of autochthonous fulvic acids from algae or phytoplankton under both photorespiration by natural sunlight or arti-ficial light, and microbial respiration or assimilation under dark incubation.
(iii) Characterization of the extracted autochthonous fulvic acids to examine the presence of functional groups, elemental composition, and possible molecular structure with reference to standard Suwannee River Fulvic Acid and Humic Acid.
(vi) Investigation on lakes having reduced DOC contents, using incorporation of ter –
restrial soils in lake surface waters. (v) Investigation on lakes having increased DOC contents, trying to reduce photosynthesis and primary production in the lake surface waters. (vi) Joint chemical and toxicological evaluation of emerging contaminants and their transformation byproducts, for important end points and target organs and effects such as mutagenicity, carcinogenicity, hepatotoxicity, nephrotoxicity, immunotoxicity, neurotoxicity, developmental neurotoxicity and pharmacokinetics (Farré et al. 2008).
Nomenclature
ABC ATP binding cassette
ADAFs Aircraft deicing/antiicing fluids
AEOs Alkylphenol ethoxylates
BF4– Tetrafluoroborate
CDOM Colored and chromopheric dissolved organic matter

90 K. M. G. Mostofa et al.
(CF 3SO2)2N– Bis(trifluoromethylsulfonyl)-imide
(CN) 2N– Dicyanamide
DBPs Disinfection byproducts (DBPs)
DDT Dichlorodiphenyltrichloroethane
DIC Dissolved inorganic carbon (DIC is defined jointly as dis
solved CO 2, H2CO3, HCO 3–, and CO 32–)
DOC Dissolved organic carbon
DOM Dissolved organic matter
DON Dissolved organic nitrogen
DOP Dissolved organic phosphorus
EDB Ethylene dibromide
EDC Endocrine-disrupting compounds
EEM Excitation-emission matrix
FDOM Fluorescent dissolved organic matter
FTOHs Fluorinated telomer alcohols
HIV Human immunodeficiency virus
H2O2 Hydrogen peroxide
IHSS International humic substances society
LMW Low molecular weight
MXR Multixenobiotic resistance
OM Organic matter
O2•− Superoxide radical
HO• Hydroxyl radical
NDMA N-nitrosodimethylamineN-EtFOSAA N-ethyl perfluorooctane sulfonamide acetate
NHDEC Neohesperidin dihydrochalcone
NPEOs Nonylphenol polyethoxylates
PAR Photosynthetically available radiation
PCBs Polychlorinated biphenyls
PF
6− Hexafluorophosphate
PFBA Perfluorobutanoic acid
PFCs Perfluorinated compounds
PFEtS Perfluoroethane sulfonate
PFHxA Perfluorohexanoic acid
PFOA Perfluorooctanoic acid
PFOS Perfluorooctane sulfonate
PFOSA Perfluorooctane sulfonamide
PFPeA Perfluoropentanoic acid
PFPrA Perfluoropropanoic acid
PFPrS Perfluoropropane sulfonate
POM Particulate organic matter
PPCPs Pharmaceuticals, personal care products
PCPs Personal care products
SRFA Suwannee River Fulvic Acid
SRHA Suwannee River Humic Acid

91 Dissolved Organic Matter in Natural Waters
THMs Trihalomethanes
TOC Total organic carbon
UV Ultraviolet1 mg L
−1 (1 × 1000)/12 = 83 μM C
Problems
(1) Define the dissolved organic matter (DOM) and explain how does it differ
from organic matter?
(2) What are the major sources of DOM in natural waters?
(3) Explain the DOM functions shortly.
(4) Explain the origin of allochthonous DOM in soil and autochthonous DOM in natural waters.
(5) What are the contributions of humic substances (fulvic and humic acids) in groundwater, rivers, lakes and oceans?
(6) Explain the redox behavior of fulvic and humic acids.
(7) Define the allochthonous fulvic and humic acids, and the autochthonous ful-vic acids. What are the chemical differences among these classes of humic substances?
(8) Why does the molecular size of DOM decrease from rivers to lakes and from lakes to oceans?
(9) What are the controlling factors that affect the DOM contents in natural waters? Explain the two most important factors that affect DOM in natural waters.
(10) Explain the possible mechanisms for the increased or declined DOM con-
tents in surface waters.
(11) What are the emerging contaminants? Explain the sources, transportation
and toxicological effects of these contaminants in the aquatic environments.
(12) How does DOM act as energy source for living organisms and aquatic
ecosystems?
Acknowledgments We are particularly grateful to Dr. Liu Cong-Qiang, Professor and
Academician; Dr. Hu Ruizhong, Professor and Director General; Dr. Wang Shijie, Professor
and Vice-director; Dr. Feng Xin Bin, Professor and Vice Director; Prof. Yun Liu, Prof. Xiao Tangfu, Dr. Li Xiao-Dong, Dr. Ding Hu of Institute of Geochemistry, Chinese Academy of Sciences; Mrs. Asma Mostofa and Mrs. Rafia Khatun for their kind assistances, constant support and inspiration during the preparations of the primary and final draft of the manuscripts. This work was financially supported by the Institute of Geochemistry, the Chinese Academy of Sciences, China. This work was partly supported by Center for Innovation and Entrepreneurship, Northwest Missouri State University, USA; Atmospheric and Ocean Research Institute, The
University of Tokyo, Japan; University of Turin and Centro Interdipartimentale NatRisk,
I-10095 Grugliasco (TO), Italy; Kyoto University, Japan; and Chinese Research Academy of Environmental Sciences, China. This chapter acknowledges Ramirez AJ, Mottaleb MA, Brooks BW, Chambliss CK, 2007, Analysis of Pharmaceuticals in Fish Using Liquid Chromatography-Tandem Mass Spectrometry, Analytical Chemistry, 79 (8), 3155–3163. Copyright (2007)
Americal Chemical Society; reprinted from Journal of Chromatography A, 1216 (5), Mottaleb
MA, Usenko S, O’Donnell JG, Ramirez AJ, Brooks BW, Chambliss CK, Gas chromatography–
mass spectrometry screening methods for select UV-filters, synthetic musks, alkylphenols, an

92 K. M. G. Mostofa et al.
antimicrobial agent, and an insect repellent in fish, 815–823, Copyright (2009), with permission
from Elsevier; and copyright (2007) by the Association for the Sciences of Limnology and Oceanography, Inc.
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