Blackening and odorization of urban rivers: a bio-geochemical process

Blackening and odorization of urban rivers: a bio-geochemical process Abstract Urban rivers constitute a major part of urban drainage systems, and play critical roles in connecting other surface waters in urban areas. Black-odorous urban rivers are widely found in developing countries experiencing rapid urbanization, and the mismatch between urbanization and sewage treatment is thought to be the reason. The phenomena of blackening and odorization are likely complex bio-geochemical processes of which the microbial interactions with the environment are not fully understood. Here, we provide an overview of the major chemical compounds, such as iron and sulfur, and their bio-geochemical conversions during blackening and odorization of urban rivers. Scenarios explaining the formation of black-odorous urban rivers are proposed. Finally, we point out knowledge gaps in mechanisms and microbial ecology that need to be addressed to better understand the development of black-odorous urban rivers. black-odorous, urban river, bio-geochemical process, blackening and odorization, sediment INTRODUCTION Rivers and lakes serve urban populations as water resources and drainage systems. They play important roles as domestic, industrial and agricultural water resources. Urban rivers are also a convenient route of transportation and as centers for aquatic recreation impact on property prices and city development decisions. However, rapid urbanization caused by fast population growth often does not keep pace with construction of sewage treatment systems, resulting in visible and smellable pollution of urban rivers. Historically, rapid urbanization has always been accompanied by urban river pollution. In London in 1855, the English scientist and inventor Michael Faraday wrote to The Times after his passage across the river Thames: ‘The smell [of the river] was very bad, and common to the whole of the water; it was the same as that which now comes up from the gully-holes in the streets; the whole river was for the time a real sewer’ (Faraday 1855). More recently, many developing countries have experienced the problem of polluted urban rivers as well. River pollution's most visible manifestation is a change in color, usually to black, often accompanied by strong unpleasant odors. In China, after the first report of blackening and odorization of the Huangpu River in 1983 (Gu and Cai 1983), similar phenomena were observed in other urban rivers as well as in their tributaries throughout the whole country, e.g. the Suzhou River in Shanghai (Ying, Zhang and Wu 1997), the Pearl River Delta, in particular in Guangzhou (Luo 2001), and the Weigong River in Shenyang (Li, Zhang and Yu 2003). By the end of 2016, there were around 1880 identified black-odorous urban rivers in 295 Chinese cities, and 64% of those rivers were located in coastal areas of southern China (Zhu 2016). The Chinese government recently released a national plan for water pollution control, the ‘Action Plan for Prevention and Treatment of Water Pollution’, and set targets for cleaning up polluted urban rivers. Nevertheless, only 45% of the identified black-odorous urban rivers were under or had finished treatment at that time (Kong 2016). Better pollution control is an urgent need to address blackening and odorization of urban rivers. However, it is often unclear what the exact reasons for the observed phenomena are. To better allocate resources for effective pollution control, it is necessary to identify the sources of pollution. A thorough understanding of the bio-geochemical processes underlying formation of black-odorous rivers is the first step, not only to apply effective pollution control but also to monitor the success of these measures and make adjustments if necessary. Organic pollutants from untreated waste streams or other non-point sources, e.g. agricultural and urban storm water runoffs (McCoy et al.2015), along river banks are believed to trigger blackening and odorization of urban rivers (Zhou, Gibson and Foy 2000; Fang et al.2012; Zhu 2016). The common understanding of the process is that high organic loading quickly depletes dissolved oxygen, leading to anaerobic conditions. Then, anaerobic microorganisms degrade dissolved organics, such as carbohydrates, fatty acids and proteins, into smaller molecules including odorous organic acids and reduced sulfur compounds, e.g. H2S and organic sulfides. These small molecules may then further react with minerals in the water and sediment, and, mediated by microorganisms, form black precipitates (Stahl 1979; Ji et al.2016). Inorganic fertilizer pollutants, such as phosphorous and nitrogen, are involved in odorization of urban rivers as well. They, for example, accelerate growth of phototrophs, a phenomenon known as eutrophication. In summer 2007, an odorous tap water crisis occurred in Wuxi, China, in which odorous volatile sulfide compounds, including methyl thiols, dimethyl sulfide and dimethyl disulfide, were produced in the river from the decomposition of massive cyanobacterial blooms (Zhang et al.2010). While such events are well documented for seawater environments (Yan, Zhou and Zou 2002), freshwater blooms have been less monitored in China. For example, the notoriously cyanobacteria-infested Lake Taihu in China experienced major blooms about every 3 years between 1960 and 1996 with increasing magnitude and frequency in recent years due to massive fertilizer pollution (Chen et al.2003). In this review, we focus on the key mechanisms and compounds involved in blackening and odorization of urban rivers. Based on the most relevant bio-geochemical processes, we propose scenarios to describe the formation of black-odorous urban rivers. We describe microbial communities in polluted and pristine freshwater systems that catalyze these processes. Lastly, we discuss challenges and possible strategies to control blackening and odorization, and propose key questions to be addressed in future studies. ELEMENTS AND COMPOUNDS CONTRIBUTING TO BLACKENING AND ODORIZATION OF URBAN RIVERS Large quantities of anthropogenic pollutants, both organic and inorganic, destabilize urban river ecosystems. The composition and concentration of organic matter in water, soil and sediment varies (Table 1). Biorecalcitrant humus, which is dominating fully decomposed organic matter, accounts for ≥40% of total organic matter present in urban rivers (Thurman 1985). These humic substances are resistant to further microbial degradation, and form black chelates with metal ions (Davies, Ghabbour and Khairy 1998; Fiedler et al.2002). Table 1. Differences between organic matter in soil and surface water. Soil  Surface water  Soil type  SOM content (%)  Humus (%)  Source  DOC (mg/L)  Humus (mg/L)  Histosols  >80  32–60  Sea water  0.2–2.0  0.06–0.6  Most mineral soils  <5  <3  River  1.0–10  0.5–4.0  Tropical soils  2  0.8–1.2  Lake  1–50  0.5–40  Soil  Surface water  Soil type  SOM content (%)  Humus (%)  Source  DOC (mg/L)  Humus (mg/L)  Histosols  >80  32–60  Sea water  0.2–2.0  0.06–0.6  Most mineral soils  <5  <3  River  1.0–10  0.5–4.0  Tropical soils  2  0.8–1.2  Lake  1–50  0.5–40  DOC, dissolved organic carbon; SOM, soil organic matter. Sources: Thurman 1985; Stanley 2000; Juo and Franzluebbers 2003; He et al.2010; Zhang et al.2012; Osman 2013; Tfaily et al.2017. View Large The abundance of inorganic substances in urban river sediments is similar to that in the Earth's crust and soil (Table 2). Dominant metallic elements in river waters are iron, magnesium, aluminum and manganese, which originate from major clay minerals in sediments (Table 3; Abdullah et al.2014). Abundant metals in the Earth's crust such as Fe and Mn are major blackening ingredients in black-odorous urban rivers (Tables 2 and 3; Metzger et al.2014). Other major metallic elements, e.g. Al, Ca, Mg and Zn, are either of white color when forming minerals or their redox potentials are too low (≤−760 mV at standard conditions) for participation in natural redox processes. Sulfur, nitrogen and carbon are the three major non-metallic elements contributing to the stench of urban rivers through formation of volatile compounds, e.g. H2S, organic sulfides, NH3, amines and short chain fatty acids (Tables 2 and 3; Ginzburg et al.1999; Bentley and Chasteen 2004; Ebil, Dursun and Dentel 2014). In coastal areas, urban rivers are often tributaries of tidal rivers with high concentrations of sulfate and magnesium (Latha and Rao 2012). Additionally, metallic elements, e.g. Fe, Mn and Mg, are mobile between sediment and water phase, sometimes mediated by microorganisms (Rzepecki 2012). These exchange rates are accelerated by organic pollution of urban rivers (Odigie et al.2014). Table 2. The content of main elements in crust, soil, surface sediment and surface water. Elements  Crust (%)  Soil (%)  Surface sediments (%)  Surface water (ppm)  O  46–50  49      Si  26–27  33      Al  7.5–8.3  7.1  6.72  0.72084  Fe  4.7–5.8  4  2.61  0.80041  Ca  3.39–5.2  1.5  1.2    K  2.3–2.64  1.4  2.46    Na  1.7–2.4  0.15  2.39    Mg  1.87–2.8  0.5  1.24  6.158  Ti  0.45–0.64  0.5  0.3186    Cl  0.13–0.19  0.01      P  0.09–0.12  0.08  0.05485    C  0.02–0.09  2      Mn  0.08–0.10  0.1  0.06239  0.01697  S  0.026–0.048  0.07      N  0.002–0.003  0.2      Cr  0.01–0.03  0.007  0.00997  0.00588  F  0.054–0.059  0.02      Ni  0.008–0.009  0.005  0.00829  0.00188  V  0.009–0.019  0.009  0.00547    Co  0.0018–0.0025  0.0008  0.00114  0.00045  Cu  0.005–0.006  0.003  0.0108  0.00626  Zn  0.007–0.009  0.0009  0.0388  0.010943  Pb  0.0012–0.0016  0.0035  0.00547  0.00807  As  0.00018–0.00022  0.0006  0.000885  0.003108  Br  0.00021–0.00025  0.001      Cd  0.000015–0.00002  3.5E-05  0.0000778  0.00007  Elements  Crust (%)  Soil (%)  Surface sediments (%)  Surface water (ppm)  O  46–50  49      Si  26–27  33      Al  7.5–8.3  7.1  6.72  0.72084  Fe  4.7–5.8  4  2.61  0.80041  Ca  3.39–5.2  1.5  1.2    K  2.3–2.64  1.4  2.46    Na  1.7–2.4  0.15  2.39    Mg  1.87–2.8  0.5  1.24  6.158  Ti  0.45–0.64  0.5  0.3186    Cl  0.13–0.19  0.01      P  0.09–0.12  0.08  0.05485    C  0.02–0.09  2      Mn  0.08–0.10  0.1  0.06239  0.01697  S  0.026–0.048  0.07      N  0.002–0.003  0.2      Cr  0.01–0.03  0.007  0.00997  0.00588  F  0.054–0.059  0.02      Ni  0.008–0.009  0.005  0.00829  0.00188  V  0.009–0.019  0.009  0.00547    Co  0.0018–0.0025  0.0008  0.00114  0.00045  Cu  0.005–0.006  0.003  0.0108  0.00626  Zn  0.007–0.009  0.0009  0.0388  0.010943  Pb  0.0012–0.0016  0.0035  0.00547  0.00807  As  0.00018–0.00022  0.0006  0.000885  0.003108  Br  0.00021–0.00025  0.001      Cd  0.000015–0.00002  3.5E-05  0.0000778  0.00007  Sources: Gaillardet, Viers and Dupré 2003; Yang et al.2003; JeffersonLab 2007; Liu et al.2007; Landaud, Helinck and Bonnarme 2008; Viers, Dupré and Gaillardet 2009; Feng et al.2010; Lin et al.2012; Song et al.2013; Gao et al.2016; Song et al.2017. View Large Table 3. Major clay minerals composition and content (%) in river sediments. Minerals  Pearl  Pearl River  Huanghe  Changjiang  Changjiang  Molecular    River  estuary  River  River  estuary  formula  Kaolinite  46  40  10  16  10  Al2Si2O5(OH)4              Clinochlore: (Mg5Al)(AlSi3)O10(OH)8              Chamosite: (Fe5Al)(AlSi3)O10(OH)8  Chlorite  25  28  16  12  26                Nimite: (Ni5Al)(AlSi3)O10(OH)8              Pennantite: (Mn,Al)6(Si,Al)4O10(OH)8  Illite  26  26  62  66  58  (K,H3O)(Al,Mg,Fe)2(Si,Al)4O10[(OH)2,(H2O)              Montmorillonite: (Na,Ca)0.33(Al,Mg)2(Si4O10)(OH)2·nH2O  Smectite  <2  <6  <12  <6  3                Nontronite: Na0.3Fe2((Si,Al)4O10)(OH)2·nH2O              Saponite: Ca0.25(Mg,Fe)3((Si,Al)4O10)(OH)2·nH2O  Minor              mineral  <1  <1  NP  NP  <3  FeS2  (Pyrite)              Minerals  Pearl  Pearl River  Huanghe  Changjiang  Changjiang  Molecular    River  estuary  River  River  estuary  formula  Kaolinite  46  40  10  16  10  Al2Si2O5(OH)4              Clinochlore: (Mg5Al)(AlSi3)O10(OH)8              Chamosite: (Fe5Al)(AlSi3)O10(OH)8  Chlorite  25  28  16  12  26                Nimite: (Ni5Al)(AlSi3)O10(OH)8              Pennantite: (Mn,Al)6(Si,Al)4O10(OH)8  Illite  26  26  62  66  58  (K,H3O)(Al,Mg,Fe)2(Si,Al)4O10[(OH)2,(H2O)              Montmorillonite: (Na,Ca)0.33(Al,Mg)2(Si4O10)(OH)2·nH2O  Smectite  <2  <6  <12  <6  3                Nontronite: Na0.3Fe2((Si,Al)4O10)(OH)2·nH2O              Saponite: Ca0.25(Mg,Fe)3((Si,Al)4O10)(OH)2·nH2O  Minor              mineral  <1  <1  NP  NP  <3  FeS2  (Pyrite)              NP, not provided. Sources: Liu et al.2007; Wang et al.2006; Yang et al.2003. View Large BIO-GEOCHEMICAL TRANSFERS OF THE KEY ELEMENTS INVOLVED IN WATER BLACKENING AND ODORIZAITON Black color formation via metal precipitates Black matter in urban rivers comprises black metallic precipitates and precipitates of brown, green, or other colors that together form a dark color. In O2-depleted surface waters, metals precipitate with sulfide and stain the water black (Fig. 1; Table 4; Nealson and Little 1997; Metzger et al.2014). Common metals such as iron and nickel form black or dark sulfides. Iron, nickel and copper sulfides are the thermodynamically most favorable precipitates (Table 4). Stahl (1979) investigated a black water lake in Illinois and demonstrated that ferrous sulfide was responsible for the black color. Mixed minerals such as copper–iron sulfides have been observed as well, for example, in the Danube River Basin (Brankov, Milijašević and Milanović 2012). Copper and other heavy metal precipitates were also detected in the Pearl River Estuary, China (Fang, Li and Zhang 2005) as well as the Reno River watershed, Italy (Ferronato et al.2013). In reduced environments, Mn exists as soluble Mn2+, due to its low affinity for sulfur, and dose not precipitate (Nealson and Little 1997). Figure 1. View largeDownload slide The microbially mediated reduction of Fe/Mn-minerals in urban rivers. Figure 1. View largeDownload slide The microbially mediated reduction of Fe/Mn-minerals in urban rivers. Table 4. Thermodynamics of some black or dark metal mineral reactions. Net reaction    ΔG°΄/M  ΔG°΄/S      (kJ mol−1)  (kJ mol−1)  Reductive environments      SO42− + H3C–COO− + 3 H+  →  HS− + 2 HCO3− + 3 H+  n/a  −48  Fe2+ + HS− + H+  →  FeSa + 2 H+  −231  −231  8 FeOOHb + 9 H3C–COO− + 8 SO42− + 25 H+  →  8 FeS + 18 HCO3− + 18 H+ + 12 H2O  −93  −93  2 FeOOH + 3 HS− + 3 H+  →  FeS + FeS2c + 4 H2O  −74  −50   Fe3O4 + 4 HS− + 4 H+  →  2 FeS + FeS2 + 4 H2O  −61  −46  FeS + S0  →  FeS2  −60  −30  4 FeOOH + 6 HS− + 6 H+  →  FeS2 + Fe3S4d + 8 H2O  −57  −38  2 FeOOH + 3 HS− + 3 H+  →  2 FeS + S0 + 4 H2O  −45  −30  Fe3O4e + 4 HS− + 4 H+  →  Fe3S4 + 4 H2O  −38  −28  9 FeS + 5 HS− + 5 H+  →  3 FeS2 + 2 Fe3S4  −2  −1  4 S0 + H3C–COO− + H+ + 4 H2O  →  4 HS− + 2 HCO3− + 6 H+  n/a  −2  24 FeOOHa + H3C–COO− + H+  →  8 Fe3O4 + 2 HCO3− + 2 H+ + 12 H2O  −5  n/a  8 FeOOH + H3C–COO− + 17 H+  →  8 Fe2+ + 2 HCO3− + 2 H+ + 12 H2O  186  n/a  Ni2+ + HS− + H+  →  NiSf + 2 H+  −184  −184  Cu2+ + HS− + H+  →  CuSg + 2 H+  −171  −171  CuS + S0  →  CuS2h  −33  −16  Pb2+ + HS− + H+  →  PbSi + 2 H+  −126  −126  Zn2+ + HS− + H+  →  ZnSj + 2 H+  −106  −106  Mn2+ + HS− + H+  →  MnSk + 2 H+  −42  −42  MnS + S0  →  MnS2l  167  167  Oxidative environments  ΔG°΄/Fe  ΔG°΄/ox    (kJ mol−1)  (kJ mol−1)  2 Cr2O3m + 3 O2 + 4 H2O  →  4 HCrO4− + 4 H+  −1094  −1459  10 FeS + 6 NO3− + 6 H+ + 2 H2O  →  10 FeOOH + 10 S0 + 3 N2  −250  −417  4 FeS + 3 O2 + 5 H2O  →  4 FeOOH + 4 S0 + 3 H2O  −270  −359  HS− + MnO2n + 3 H+  →  S0 + Mn2+ + 2 H2O  n/a  −130  2 FeS + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ 2 S0 + 2 H2O  −150  −100  2 FeS2 + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ + 4 S0 + 2 H2O  −90  −60  Net reaction    ΔG°΄/M  ΔG°΄/S      (kJ mol−1)  (kJ mol−1)  Reductive environments      SO42− + H3C–COO− + 3 H+  →  HS− + 2 HCO3− + 3 H+  n/a  −48  Fe2+ + HS− + H+  →  FeSa + 2 H+  −231  −231  8 FeOOHb + 9 H3C–COO− + 8 SO42− + 25 H+  →  8 FeS + 18 HCO3− + 18 H+ + 12 H2O  −93  −93  2 FeOOH + 3 HS− + 3 H+  →  FeS + FeS2c + 4 H2O  −74  −50   Fe3O4 + 4 HS− + 4 H+  →  2 FeS + FeS2 + 4 H2O  −61  −46  FeS + S0  →  FeS2  −60  −30  4 FeOOH + 6 HS− + 6 H+  →  FeS2 + Fe3S4d + 8 H2O  −57  −38  2 FeOOH + 3 HS− + 3 H+  →  2 FeS + S0 + 4 H2O  −45  −30  Fe3O4e + 4 HS− + 4 H+  →  Fe3S4 + 4 H2O  −38  −28  9 FeS + 5 HS− + 5 H+  →  3 FeS2 + 2 Fe3S4  −2  −1  4 S0 + H3C–COO− + H+ + 4 H2O  →  4 HS− + 2 HCO3− + 6 H+  n/a  −2  24 FeOOHa + H3C–COO− + H+  →  8 Fe3O4 + 2 HCO3− + 2 H+ + 12 H2O  −5  n/a  8 FeOOH + H3C–COO− + 17 H+  →  8 Fe2+ + 2 HCO3− + 2 H+ + 12 H2O  186  n/a  Ni2+ + HS− + H+  →  NiSf + 2 H+  −184  −184  Cu2+ + HS− + H+  →  CuSg + 2 H+  −171  −171  CuS + S0  →  CuS2h  −33  −16  Pb2+ + HS− + H+  →  PbSi + 2 H+  −126  −126  Zn2+ + HS− + H+  →  ZnSj + 2 H+  −106  −106  Mn2+ + HS− + H+  →  MnSk + 2 H+  −42  −42  MnS + S0  →  MnS2l  167  167  Oxidative environments  ΔG°΄/Fe  ΔG°΄/ox    (kJ mol−1)  (kJ mol−1)  2 Cr2O3m + 3 O2 + 4 H2O  →  4 HCrO4− + 4 H+  −1094  −1459  10 FeS + 6 NO3− + 6 H+ + 2 H2O  →  10 FeOOH + 10 S0 + 3 N2  −250  −417  4 FeS + 3 O2 + 5 H2O  →  4 FeOOH + 4 S0 + 3 H2O  −270  −359  HS− + MnO2n + 3 H+  →  S0 + Mn2+ + 2 H2O  n/a  −130  2 FeS + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ 2 S0 + 2 H2O  −150  −100  2 FeS2 + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ + 4 S0 + 2 H2O  −90  −60  Black or dark minerals: airon sulfide, bgoethite, cpyrite, dgreigite, emagnetite, fmillerite, gcovellite, hα-chalkosite, ilead sulfide, jsphalerite (disulfide not known for Zn and Pb), kalabandite (pink, orange or green), lhauerite, meskolaite, npyrolusite (light grey). M, metal; n/a, not applicable; ox. oxidant. View Large While iron sulfide formation is spontaneous, microorganisms such as Geobacter, Geothrix, Rhodoferax and Shewanella can harvest the released energy for their cell growth (Fig. 1; Lovley 1991). Thermodynamically, formation of FeS is favored followed by FeS2 and Fe3S4 (Table 4). Greigite (Fe3S4) is formed in excess of sulfide. Elemental sulfur (S0) and polysulfide (Sn2−) formation are thought to be the intermediate steps leading to pyrite and greigite (Table 4; Rickard 1975; Luther 1991). Similar to bio-geochemical transfers in marine sediments, the FeS and FeS2 as well as other metals play central roles in the sulfur cycle in urban rivers (Schippers and Jørgensen 2002). That is, metals and especially abundant iron, are reduced by organic pollutants, such as volatile fatty acids (such as acetic, butyric and propionic acids), or other reducing equivalents (Table 4). If iron then again enters oxidizing zones, for example by currents or shift of oxic zones, it can be re-oxidized by dissolved oxygen, nitrate, or manganese oxides. Iron oxides, such as goethite, can also be reduced biologically. Humic substances enhance the bioavailability of insoluble Fe(III) oxides as electron acceptors and therefore improve the thermodynamics of biological iron reduction (Lovley et al.1996, 1998). Quinone moieties in humic substances serve as electron shuttles in Fe(III)-respiring microorganisms, e.g. Ferribacterium limneticum and Geobacter metallireducens, accelerating the rate of both Fe(III) oxide reduction in river sediments and contaminant oxidation coupled to Fe(III) reduction (Lovley et al.1996; Finneran and Lovley 2001; Nevin and Lovley 2000). Other iron reducers such as Shewanella oneidensis (Venkateswaran et al.1999), Paludibaculum fermentans (Kulichevskaya et al.2014) and Anaeromyxobacter dehalogenans (Sanford, Cole and Tiedje 2002) are able to utilize a number of different electron donors including sugars and long chain fatty acids. The broad variety of electron acceptors and donors used by iron reducers makes these microorganisms ubiquitous in freshwater sediments. In sediments, pyrite (FeS2) is oxidized abiotically at mineral interfaces, for example between FeS2 and MnO2 (Table 4; Schippers and Jørgensen 2002). Immediate products of this oxidation are thiosulfate and polythionates, which can be further oxidized to sulfate by manganese-reducing bacteria (Jørgensen and Nelson 2004). Additionally, Fe(II) and Mn(II) were released from sediment pore waters to form black oxides in anoxic–oxic water interface zone of the black rivers (Atkinson et al.2007). Small amounts of dissolved ferrous iron released from FeS2 in sediments can be re-oxidized by oxygen, nitrate and MnO2 and precipitated as black magnetite in river beds. After reentry in oxic zones or through mediation, reduced Mn(II) is recycled to MnO2 via microbial oxidation in the presence of, even trace amounts of, O2 or nitrate in surface waters (Boogerd and de Vrind 1987; Marcus et al.2017). Nitrate has therefore been suggested as a cost-effective remediation method for black urban rivers (He et al.2017). Odorous volatile compounds Sulfur compounds Odorous compounds in urban rivers are volatile organic and inorganic compounds. Volatile sulfur compounds generated from microbial sulfate reduction or degradation of sulfur-containing organic matter normally have an unpleasant odor, including that of inorganic H2S and organic sulfides (Kadota and Ishida 2003). Sources of such reduced sulfur compounds vary. River deltas discharging into oceans often experience seawater influx due to tidal activity. Therefore, sulfide, as a result of microbial sulfate reduction, is detected in significant amounts in urban rivers connecting to major deltas in coastal areas, particularly in the Pearl River estuary where 3 mM sulfide in the sediment was reported as a result of sulfate reduction (Fang et al.2005). Most sulfide in estuaries is produced by sulfate reducing microorganisms using dissimilatory sulfate reduction for respiration. Sulfate reducing microorganisms are broadly dispersed across the prokaryotic phylogenetic tree but are often found among δ-Proteobacteria and Firmicutes as well as Archaea (Zhou et al.2011). In addition to the production of odorous H2S, sulfate reduction may eventually lead to the formation of black iron sulfide species even with trace amounts of iron in the water and sediment (Wu et al.2016). Another source of sulfide in urban rivers is organic sulfur (Giordano et al.2005). It enters the sulfur cycle via assimilatory sulfate reduction (Fig. 2). The amount of sulfur in domestic waste streams or other anthropogenic sources, however, is negligible compared with the massive cyanobacterial blooms in the surface waters (Zhang et al.2010). CO2-fixing cyanobacteria frequently are the main source of organic matter in surface waters. They assimilate sulfur via cysteine/methionine biosynthetic pathways (Fig. 2). This organic sulfur is subsequently released in the form of volatile organic sulfur compounds, which include thiols and thioethers in methylated sulfide species, e.g. methyl sulfide, dimethyl sulfide and dimethyl disulfide, as byproducts—all of which are characterized by their notoriously bad smell (Bentley and Chasteen 2004). Figure 2. View largeDownload slide Methionine biosynthesis (top) and degradation (bottom) proceed via cysteine. The biosynthetic pathway requires ATP for sulfate reduction to sulfide, producing the intermediate 3΄-phosphoadenosine-5΄-phosphosulfate (PAPS). Together with serine, sulfide then forms cysteine and ultimately methionine. Under anaerobic conditions, methionine is degraded to taurine, CO2 and sulfide in exchange for sulfite. Figure 2. View largeDownload slide Methionine biosynthesis (top) and degradation (bottom) proceed via cysteine. The biosynthetic pathway requires ATP for sulfate reduction to sulfide, producing the intermediate 3΄-phosphoadenosine-5΄-phosphosulfate (PAPS). Together with serine, sulfide then forms cysteine and ultimately methionine. Under anaerobic conditions, methionine is degraded to taurine, CO2 and sulfide in exchange for sulfite. Nitrogen and organic carbon compounds Nitrogen compounds are the largest group of malodorous compounds generated by proteolytic microorganisms, and their smells range from that of ammonia to the typical smell of corpse decomposition (Wang et al.2017). Microbiogenic malodorous nitrogen compounds include organic amines such as cadaverine (1,5-pentanediamine) and putrescine (1,4-butanediamine). Cadaverine is produced via decarboxylation of lysine, whereas putrescine is a product of degradation of ornithine, an essential building block of bacterial cell walls (Wunderlichová et al.2014; Ma et al.2017). Other abundant malodorous nitrogen compounds are volatile alkylated amines of characteristic fishy smell, which is sensed even in trace amounts. These compounds comprise alkylated amines, such as methylamine, dimethylamine, trimethylamine, ethaneamine, propaneamine and butaneamine (Ge et al.2011). Methylamines are degradation products of N-methylated amino acids, with glycine, betaine, choline, trimethylamine and carnitine as their natural precursors in biomass, which, in turn, is introduced into urban rivers via wastewater or algal blooms (Ikawa and Taylor 1973). All organisms are able to hydrolyze proteins using proteases as this is an essential part of their metabolism. Cell internal proteolysis is necessary in every organism, for example to tune its enzymatic machinery to novel environmental conditions or to control vital cell functions. Microorganisms specifically feeding on peptides can be isolated using casamino acids and trypticase peptone media, and often yield strains closely related to Clostridium species when grown anaerobically. Such proteolytic microorganisms are ubiquitous in anaerobic and aerobic environments alike. Examples of anaerobic environments are rumens (Blackburn and Hobson 1962), anaerobic digesters (Abendroth et al.2015), peat bogs (Juottonen et al.2005) and rice paddies (Weber et al.2001). Typical aerobic environments are many processed food products such as cabbage (Borla, Davidovich and Roura 2010) and dairy products (Frazier and Rupp 1931). Despite the presence of proteolytic microorganisms, none of the mentioned environments are known for their obnoxious smell. The reason is that protein concentrations are either relatively low or, in the case of food products, aerobic conditions prevail. When oxygen is absent, alkylated amines cannot be further oxidized and serve as substrates for sulfate reducers or methanogens (Lovley and Klug 1983). However, since alkylated amines are gaseous or at least volatile, they often escape before slow growing anaerobic microorganisms are able to degrade them, causing the typical smell in surface waters. Odorous organic compounds without S and N elements are mostly VFAs, which are generated by anaerobic fermentation of organic pollutants in the waste streams or of decomposed compounds produced by algal blooms (Verstraete et al.1996; Pham et al.2012). VFAs in urban rivers play critical roles in coupling organic carbon compounds with iron and sulfur cycles in the surface water (Fig. 4). For example, when sulfate is present, VFAs can be further used as electron donors by sulfate reducing microorganisms and produce malodorous sulfide (Hao et al.2009). Figure 3. View largeDownload slide Principal coordinate analysis (PCoA with Bray–Curtis distance matrices) of six surface-water sediment samples collected from three different geographic areas, i.e. Pearl River Delta, Yangtze River Delta and Dongting area. This figure is plotted with published 16S rRNA gene-sequencing data (Wang et al2012; Liu et al2014; He et al2017; Huang et al2017). The trophic states of the six surface waters are: eutrophic for Taihu Lake and Chebei River (black-odorous river); mesotrophic for Dongting Lake, Huangda Lake and Pearl River; and oligotrophic for Liuxi Reservoir. Figure 3. View largeDownload slide Principal coordinate analysis (PCoA with Bray–Curtis distance matrices) of six surface-water sediment samples collected from three different geographic areas, i.e. Pearl River Delta, Yangtze River Delta and Dongting area. This figure is plotted with published 16S rRNA gene-sequencing data (Wang et al2012; Liu et al2014; He et al2017; Huang et al2017). The trophic states of the six surface waters are: eutrophic for Taihu Lake and Chebei River (black-odorous river); mesotrophic for Dongting Lake, Huangda Lake and Pearl River; and oligotrophic for Liuxi Reservoir. MICROBIAL ECOLOGY OF PRISTINE AND POLLUTED FRESHWATER ENVIRONMENTS Urban river pollution affects microbial communities in water and sediments with measurable effects in the short (Schöll and Szövényi 2011) and long term (Ibekwe, Ma and Murinda 2016; Lu, Chen and Zheng 2017). This makes microbial community analysis an additional monitoring tool for water quality (Drury, Rosi-Marshall and Kelly 2013; García-Armisen et al.2014; Li, Sharp and Drewes 2016; Xie et al.2016; Köchling et al.2017). Diversity (Drury, Rosi-Marshall and Kelly 2013; Staley et al.2013), richness (Lin et al.2014) and variability (Lu, Chen and Zheng 2017) of microbial communities have been affected by anthropogenic pollutants. As expected, coliform growth is correlated with fecal anthropogenic contamination, for example in large rivers such as the Danube (Hoch et al.1996; Kirschner et al.2009) and the Mississippi (Staley et al.2013), as well as smaller rivers such as the Jaboatão River in Brazil (Köchling et al.2017), the Reno River in Italy (Ferronato et al.2013) and small creeks of a California watershed (Ibekwe, Ma and Murinda 2016). A clear impact of treated wastewater on community composition and metabolism was reported for the Taif River in Saudi Arabia, where pristine samples showed a higher representation of carbohydrate metabolic genes along with fatty acid and amino acid anabolic genes as opposed to samples impacted by wastewater (Li, Sharp and Drewes 2016). The latter were enriched in genes associated with nitrogen and sulfur metabolism, as would be expected in nutrient rich environments. Inverse metabolic patterns were reported for river sediments in China where energy-, carbohydrate- and amino acid-related genes were enriched or equal to pristine control sediments (Lu, Chen and Zheng 2017). Algal growth is often considered to be linked to anthropogenic contamination in freshwater systems such as Taihu Lake in China (Huang et al.2017) or the Zenne River in Belgium (García-Armisen et al.2014). However, the mechanisms by which pollution and algal growth are connected are not always clear. For example, Huang et al. (2017) found that phosphate as well as organic matter concentration were correlated with cyanobacterial growth in the Taihu Lake, whereas in the Danube River, Kirschner et al. (2009) identified only a link to organic matter but not any other of the factors investigated, such as phosphate, nitrogen and temperature. This suggests that cyanobacteria live heterotrophically or that low bioavailability of inorganic nutrients are responsible for algal blooms. Despite the ongoing DNA sequencing revolution, it is not clear which factors shape river sediment communities. Some recent attempts indicate that, indeed, organic matter released into urban rivers by sewage streams promotes growth of certain microbial lineages such as Acinetobacter, Flaviobacterium, Thauera and Rhodococcus in the Zenne River flowing through the Brussels metropolitan area (García-Armisen et al.2014). A similar correlation between organic matter and Cyanobacteria was linked to fecal coliforms and Enterococci in the Danube (Kirschner et al.2009). The dependence of fecal coliforms on environmental factors, however, was stronger in the water column than in the sediments studied in selected creeks of a southern California watershed (Ibekwe, Ma and Murinda 2016). In the Rhône River prodelta, microbial variation could be explained by organic matter as well (Fagervold, et al.2014). In addition to organic matter, Ji et al. (2016) found that iron and sulfate concentrations as well as pH were associated with methanogenic networks identified in Amazonian lake sediments. The pioneering works of Zwart et al. (2002) and Newton et al. (2011) identified an appreciable bacterial diversity in freshwater systems showing that river and lake communities are similar. In addition, our own comparison of published 16S rRNA gene-sequencing data of three Chinese lakes, two rivers and one reservoir from three distant areas shows that geographical location best explained the differences between the investigated freshwater environments (Fig. 3). Though microbial populations of contaminated and pristine sites of the same region cluster closely, communities still show some difference. Proteobacteria are the largest phylum in the prokaryotic tree of life and are therefore also highly abundant in river sediments—polluted or not. Nearly all river and lake sediments surveyed here harbored α-, β- and γ-Proteobacteria. The most prominent representative of freshwater α-Proteobacteria is the SAR11 clade (Pelagibacter; Salcher et al.2011; Savio et al.2015). Also Bacteroidetes were found in nearly all freshwater sediments. Together, these four groups cover 40% of all cultured prokaryotic species, making their dominance in freshwater sediments only natural. Consequently, microbial communities in freshwater sediments are often very similar at the phylum level (Ji et al.2016). Nonetheless, a study screening 68 publications of lake microbial communities using only high quality Sanger-sequencing data reported a large heterogeneity at lower taxonomic levels, termed tribes (Newton et al.2011). This finding was confirmed for Mississippi River sediments where only 12% of the identified operational taxonomic units (>97% sequence identity) were shared across all sites (Staley et al.2013). It is hence the less abundant groups, such as Acidobacteria, Actinobacteria, Verrucomicrobia, Chloroflexi, Planctomycetes, Gemmatimonadetes and Archaea, or tribes that may act as distinctive indicators for metabolic processes. Many such tribes are uncultured representatives of freshwater environments and do not match with Linnaean taxonomic boundaries. While broad surveys of our drinking water resources need to be continued, it remains unclear how under-represented parts of microbial communities adapt to pollution. Indeed, a recent metagenomic survey of a wastewater impacted river showed that small community factions are major hubs in microbial assemblages (Li, Sharp and Drewes 2016). This is an observation that has also been made in the pristine Lake Cadagno, Switzerland, where 0.3% of the cells in the lake were responsible for 40% of the substrate turnover (Musat et al.2008). In conclusion, more studies on natural environments are necessary to understand the above mentioned discrepancies and to establish a baseline for future research on pollution affected environments. A SCENARIO OF THE BIO-GEOCHEMICAL PROCESS IN BLACKENING AND ODORIZAITON OF URBAN RIVERS In the scenario depicted in Fig. 4, organic matter originates either directly from anthropogenic sources, e.g. waste streams and other non-point source pollution, or from decomposition of cyanobacteria biomass. These are the main suspects in generating odors by production of volatile (organic) sulfides, odorous amines and VFAs (Van Neste et al.1987; Ginzburg et al.1998; Hu et al.2007; Zhang et al.2010). Organic matter is also the major source of reducing equivalents for the reduction of sulfate and iron to produce black minerals such as iron sulfides, which link the sulfur and iron cycles in urban rivers (Fig. 4; Berner et al.1985; Lovley 1987). Therefore, input of organic matter into urban rivers is likely the key factor for triggering water blackening and odorization. Sulfur input from sediment, seawater, or decomposing algal biomass is directly involved in the formation of black- and odorous-matter in urban rivers, e.g. via formation of sulfide species such as H2S and iron sulfides. Alkylated sulfides are often byproducts of cyanobacterial metabolism and biomass degradation and their volatility makes them strong odorous constituents of some urban rivers’ stench. Sulfate reduction by diverse sulfate reducing microorganisms dominates in sediments because their redox potential confers growth advantages to these microorganisms over their competitors. For example, the standard redox potential of sulfate reduction (−217 mV) is slightly more positive than that of hydrogenotropic methanogenic process (−240 mV). This and the higher energy gain compared with iron reduction suggest that sulfur is the link between the different element cycles as shown for Black Sea sediments (Siegert et al.2013). Figure 4. View largeDownload slide Scenarios describing bio-geochemical transfers of major black and odorous elements/compounds in urban rivers. In the process of organic matter degradation, odors result from production of volatile organic sulfur compounds, methylamines and VFAs, which escape into the atmosphere. In addition, organic matter supplies reducing power for SO42− and Fe(III) to form FeSx. The processes for black/odorous compounds are shown with solid lines, and others with dotted lines. DMS, dimethyl sulfide; VFAs, volatile organic sulfur compounds; VOSCs, volatile organic sulfur compounds. Figure 4. View largeDownload slide Scenarios describing bio-geochemical transfers of major black and odorous elements/compounds in urban rivers. In the process of organic matter degradation, odors result from production of volatile organic sulfur compounds, methylamines and VFAs, which escape into the atmosphere. In addition, organic matter supplies reducing power for SO42− and Fe(III) to form FeSx. The processes for black/odorous compounds are shown with solid lines, and others with dotted lines. DMS, dimethyl sulfide; VFAs, volatile organic sulfur compounds; VOSCs, volatile organic sulfur compounds. CONCLUSIONS AND FUTURE PERSPECTIVES The blackening and odorization of urban rivers is a complex bio-geochemical process involving five key elements, i.e. Fe, Mn, S, N and C. Outstanding issues include the following. (i) While we propose several mechanisms that contribute to blackening and odorization of urban rivers such as organic matter degradation and metal precipitation, there is no evidence yet that these are indeed the driving factors. In the past, measures to counter river pollution were taken, such as widespread treatment of industrial wastewater in the Pearl River Delta and sediment removal, yet they only mitigated the problem for a short time. Evidence for our hypothesized mechanisms needs to be collected in order to take targeted action. Gathering this evidence requires application of standard tests to assess water quality (listed in Table 5) along with novel molecular techniques and may involve the development of new methods that are more efficient. As shown in this review, the suspected blackening elements (metal sulfides) and three odor-forming elements (S, N and C) should be first targets for water quality analysis investigating blackening and odor formation in the urban rivers. Table 5. Routine methods for surface water quality assessment. Analyte  Method  References  Dissolved inorganic carbon  Spectrophotometry/potentiometry/conductimetry  Oshima et al.2001; Carlson 1978; Linares et al.1989  Dissolved organic carbon  Chemical oxidation  Sharp 1973    Ultraviolet oxidation  Beattie et al.1961; Armstrong et al.1966    High-temperature combustion  Sharp 1973; Salonen 1979  Chemical oxygen demand  Dichromate oxidation method  Moore et al.1949; Jirka and Carter 1975    Potassium permanganate oxidation method  Korenaga 1980  Metal content (Fe, Mn, Cu, Zn, etc.)  Spectroscopic analysis method (inductively coupled plasma mass spectrometry)  Houk et al. 1989  Total dissolved nitrogen  Alkaline persulfate digestion  Solorzano and Sharp 1980    High temperature oxidation  Suzuki et al.1985  Dissolved inorganic nitrogen  Phenol hypochlorite reaction method (NH3/NH4+)  Bolleter et al.1961    Nessler's reagent spectrophotometry (NH3/NH4+)  Vanselow 1940; Leonard 1963    Ion chromatography (NO2−, NO3−)  Gjerde et al.1979    Colorimetry (NO2−)  APHA 1998    Ultraviolet spectrophotometry (NO3−)  Hoather and Rackham 1959  Dissolved organic nitrogen  High-temperature catalytic oxidation  Badr et al.2003  Dissolved and precipitated sulfides  CuS colloidal solution method  Cord-Ruwisch 1985  Dissolved sulfate  Turbidimetry  Tabatabai 1974  Volatile organic sulfur compounds  Chromatography analysis method (gas chromatography–sulfur chemiluminescence detection)  Sun et al.2014    Chromatography analysis method (gas chromatography–mass spectrometry)  Van Langenhove et al.1985  Element content (C, H, N, S)  Elemental analysis  Kirsten 1971; Fadeeva et al.2008  Analyte  Method  References  Dissolved inorganic carbon  Spectrophotometry/potentiometry/conductimetry  Oshima et al.2001; Carlson 1978; Linares et al.1989  Dissolved organic carbon  Chemical oxidation  Sharp 1973    Ultraviolet oxidation  Beattie et al.1961; Armstrong et al.1966    High-temperature combustion  Sharp 1973; Salonen 1979  Chemical oxygen demand  Dichromate oxidation method  Moore et al.1949; Jirka and Carter 1975    Potassium permanganate oxidation method  Korenaga 1980  Metal content (Fe, Mn, Cu, Zn, etc.)  Spectroscopic analysis method (inductively coupled plasma mass spectrometry)  Houk et al. 1989  Total dissolved nitrogen  Alkaline persulfate digestion  Solorzano and Sharp 1980    High temperature oxidation  Suzuki et al.1985  Dissolved inorganic nitrogen  Phenol hypochlorite reaction method (NH3/NH4+)  Bolleter et al.1961    Nessler's reagent spectrophotometry (NH3/NH4+)  Vanselow 1940; Leonard 1963    Ion chromatography (NO2−, NO3−)  Gjerde et al.1979    Colorimetry (NO2−)  APHA 1998    Ultraviolet spectrophotometry (NO3−)  Hoather and Rackham 1959  Dissolved organic nitrogen  High-temperature catalytic oxidation  Badr et al.2003  Dissolved and precipitated sulfides  CuS colloidal solution method  Cord-Ruwisch 1985  Dissolved sulfate  Turbidimetry  Tabatabai 1974  Volatile organic sulfur compounds  Chromatography analysis method (gas chromatography–sulfur chemiluminescence detection)  Sun et al.2014    Chromatography analysis method (gas chromatography–mass spectrometry)  Van Langenhove et al.1985  Element content (C, H, N, S)  Elemental analysis  Kirsten 1971; Fadeeva et al.2008  View Large (ii) What, if any, are the core microbial communities, taxonomically and physiologically, mediating bio-geochemical transfers of the key elements in black-odorous urban rivers? Despite the progress in studies of metabolism and element cycles in surface waters, many puzzles remain, for example, the discrepancy in some reports and our own investigations showing that pollution sometimes does not affect microbial communities. There is a need for better understanding of the core communities coupling all these element cycles in black-odorous urban rivers. Current meta-omic technologies may help to provide in-depth insights. (iii) What is the role of minor elements, e.g. Cu and Zn, in the blackening and odorization of urban rivers? Hitherto, very few studies have investigated their contribution to blackening and odorization in surface waters. Whether these trace elements play critical roles in connecting the Fe, Mn, S, N and C cycles warrants future investigation. (iv) New water quality standards to address blackening and odorous surface waters need to be developed in China. Current guidelines are insufficient, mostly because it is not clear what the reasons for blackening and odorization are. Defining baselines will be essential to develop standards. Understanding the pathways involved in blackening and malodor-generating metabolism is key to controlling these processes and developing environmentally friendly microbial technologies. For example, scaling bio-electrochemical technologies to use polluted rivers for power production can be an environmentally friendly alternative to current treatment strategies (Ewing et al.2014). Also, microbial inhibitors may be used to block microbial participation in Fe, Mn, S, N and C cycles. FUNDING This study was supported by the Key Program of National Natural Science Foundation of China (51638005). Conflict of interest. None declared. REFERENCES Abdullah JA, Michel H, Funel GB et al.   Distribution and baseline values of trace elements in the sediment of Var River catchment. Environ Monit Assess  2014; 186: 8175– 89. 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Blackening and odorization of urban rivers: a bio-geochemical process

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Abstract

Abstract Urban rivers constitute a major part of urban drainage systems, and play critical roles in connecting other surface waters in urban areas. Black-odorous urban rivers are widely found in developing countries experiencing rapid urbanization, and the mismatch between urbanization and sewage treatment is thought to be the reason. The phenomena of blackening and odorization are likely complex bio-geochemical processes of which the microbial interactions with the environment are not fully understood. Here, we provide an overview of the major chemical compounds, such as iron and sulfur, and their bio-geochemical conversions during blackening and odorization of urban rivers. Scenarios explaining the formation of black-odorous urban rivers are proposed. Finally, we point out knowledge gaps in mechanisms and microbial ecology that need to be addressed to better understand the development of black-odorous urban rivers. black-odorous, urban river, bio-geochemical process, blackening and odorization, sediment INTRODUCTION Rivers and lakes serve urban populations as water resources and drainage systems. They play important roles as domestic, industrial and agricultural water resources. Urban rivers are also a convenient route of transportation and as centers for aquatic recreation impact on property prices and city development decisions. However, rapid urbanization caused by fast population growth often does not keep pace with construction of sewage treatment systems, resulting in visible and smellable pollution of urban rivers. Historically, rapid urbanization has always been accompanied by urban river pollution. In London in 1855, the English scientist and inventor Michael Faraday wrote to The Times after his passage across the river Thames: ‘The smell [of the river] was very bad, and common to the whole of the water; it was the same as that which now comes up from the gully-holes in the streets; the whole river was for the time a real sewer’ (Faraday 1855). More recently, many developing countries have experienced the problem of polluted urban rivers as well. River pollution's most visible manifestation is a change in color, usually to black, often accompanied by strong unpleasant odors. In China, after the first report of blackening and odorization of the Huangpu River in 1983 (Gu and Cai 1983), similar phenomena were observed in other urban rivers as well as in their tributaries throughout the whole country, e.g. the Suzhou River in Shanghai (Ying, Zhang and Wu 1997), the Pearl River Delta, in particular in Guangzhou (Luo 2001), and the Weigong River in Shenyang (Li, Zhang and Yu 2003). By the end of 2016, there were around 1880 identified black-odorous urban rivers in 295 Chinese cities, and 64% of those rivers were located in coastal areas of southern China (Zhu 2016). The Chinese government recently released a national plan for water pollution control, the ‘Action Plan for Prevention and Treatment of Water Pollution’, and set targets for cleaning up polluted urban rivers. Nevertheless, only 45% of the identified black-odorous urban rivers were under or had finished treatment at that time (Kong 2016). Better pollution control is an urgent need to address blackening and odorization of urban rivers. However, it is often unclear what the exact reasons for the observed phenomena are. To better allocate resources for effective pollution control, it is necessary to identify the sources of pollution. A thorough understanding of the bio-geochemical processes underlying formation of black-odorous rivers is the first step, not only to apply effective pollution control but also to monitor the success of these measures and make adjustments if necessary. Organic pollutants from untreated waste streams or other non-point sources, e.g. agricultural and urban storm water runoffs (McCoy et al.2015), along river banks are believed to trigger blackening and odorization of urban rivers (Zhou, Gibson and Foy 2000; Fang et al.2012; Zhu 2016). The common understanding of the process is that high organic loading quickly depletes dissolved oxygen, leading to anaerobic conditions. Then, anaerobic microorganisms degrade dissolved organics, such as carbohydrates, fatty acids and proteins, into smaller molecules including odorous organic acids and reduced sulfur compounds, e.g. H2S and organic sulfides. These small molecules may then further react with minerals in the water and sediment, and, mediated by microorganisms, form black precipitates (Stahl 1979; Ji et al.2016). Inorganic fertilizer pollutants, such as phosphorous and nitrogen, are involved in odorization of urban rivers as well. They, for example, accelerate growth of phototrophs, a phenomenon known as eutrophication. In summer 2007, an odorous tap water crisis occurred in Wuxi, China, in which odorous volatile sulfide compounds, including methyl thiols, dimethyl sulfide and dimethyl disulfide, were produced in the river from the decomposition of massive cyanobacterial blooms (Zhang et al.2010). While such events are well documented for seawater environments (Yan, Zhou and Zou 2002), freshwater blooms have been less monitored in China. For example, the notoriously cyanobacteria-infested Lake Taihu in China experienced major blooms about every 3 years between 1960 and 1996 with increasing magnitude and frequency in recent years due to massive fertilizer pollution (Chen et al.2003). In this review, we focus on the key mechanisms and compounds involved in blackening and odorization of urban rivers. Based on the most relevant bio-geochemical processes, we propose scenarios to describe the formation of black-odorous urban rivers. We describe microbial communities in polluted and pristine freshwater systems that catalyze these processes. Lastly, we discuss challenges and possible strategies to control blackening and odorization, and propose key questions to be addressed in future studies. ELEMENTS AND COMPOUNDS CONTRIBUTING TO BLACKENING AND ODORIZATION OF URBAN RIVERS Large quantities of anthropogenic pollutants, both organic and inorganic, destabilize urban river ecosystems. The composition and concentration of organic matter in water, soil and sediment varies (Table 1). Biorecalcitrant humus, which is dominating fully decomposed organic matter, accounts for ≥40% of total organic matter present in urban rivers (Thurman 1985). These humic substances are resistant to further microbial degradation, and form black chelates with metal ions (Davies, Ghabbour and Khairy 1998; Fiedler et al.2002). Table 1. Differences between organic matter in soil and surface water. Soil  Surface water  Soil type  SOM content (%)  Humus (%)  Source  DOC (mg/L)  Humus (mg/L)  Histosols  >80  32–60  Sea water  0.2–2.0  0.06–0.6  Most mineral soils  <5  <3  River  1.0–10  0.5–4.0  Tropical soils  2  0.8–1.2  Lake  1–50  0.5–40  Soil  Surface water  Soil type  SOM content (%)  Humus (%)  Source  DOC (mg/L)  Humus (mg/L)  Histosols  >80  32–60  Sea water  0.2–2.0  0.06–0.6  Most mineral soils  <5  <3  River  1.0–10  0.5–4.0  Tropical soils  2  0.8–1.2  Lake  1–50  0.5–40  DOC, dissolved organic carbon; SOM, soil organic matter. Sources: Thurman 1985; Stanley 2000; Juo and Franzluebbers 2003; He et al.2010; Zhang et al.2012; Osman 2013; Tfaily et al.2017. View Large The abundance of inorganic substances in urban river sediments is similar to that in the Earth's crust and soil (Table 2). Dominant metallic elements in river waters are iron, magnesium, aluminum and manganese, which originate from major clay minerals in sediments (Table 3; Abdullah et al.2014). Abundant metals in the Earth's crust such as Fe and Mn are major blackening ingredients in black-odorous urban rivers (Tables 2 and 3; Metzger et al.2014). Other major metallic elements, e.g. Al, Ca, Mg and Zn, are either of white color when forming minerals or their redox potentials are too low (≤−760 mV at standard conditions) for participation in natural redox processes. Sulfur, nitrogen and carbon are the three major non-metallic elements contributing to the stench of urban rivers through formation of volatile compounds, e.g. H2S, organic sulfides, NH3, amines and short chain fatty acids (Tables 2 and 3; Ginzburg et al.1999; Bentley and Chasteen 2004; Ebil, Dursun and Dentel 2014). In coastal areas, urban rivers are often tributaries of tidal rivers with high concentrations of sulfate and magnesium (Latha and Rao 2012). Additionally, metallic elements, e.g. Fe, Mn and Mg, are mobile between sediment and water phase, sometimes mediated by microorganisms (Rzepecki 2012). These exchange rates are accelerated by organic pollution of urban rivers (Odigie et al.2014). Table 2. The content of main elements in crust, soil, surface sediment and surface water. Elements  Crust (%)  Soil (%)  Surface sediments (%)  Surface water (ppm)  O  46–50  49      Si  26–27  33      Al  7.5–8.3  7.1  6.72  0.72084  Fe  4.7–5.8  4  2.61  0.80041  Ca  3.39–5.2  1.5  1.2    K  2.3–2.64  1.4  2.46    Na  1.7–2.4  0.15  2.39    Mg  1.87–2.8  0.5  1.24  6.158  Ti  0.45–0.64  0.5  0.3186    Cl  0.13–0.19  0.01      P  0.09–0.12  0.08  0.05485    C  0.02–0.09  2      Mn  0.08–0.10  0.1  0.06239  0.01697  S  0.026–0.048  0.07      N  0.002–0.003  0.2      Cr  0.01–0.03  0.007  0.00997  0.00588  F  0.054–0.059  0.02      Ni  0.008–0.009  0.005  0.00829  0.00188  V  0.009–0.019  0.009  0.00547    Co  0.0018–0.0025  0.0008  0.00114  0.00045  Cu  0.005–0.006  0.003  0.0108  0.00626  Zn  0.007–0.009  0.0009  0.0388  0.010943  Pb  0.0012–0.0016  0.0035  0.00547  0.00807  As  0.00018–0.00022  0.0006  0.000885  0.003108  Br  0.00021–0.00025  0.001      Cd  0.000015–0.00002  3.5E-05  0.0000778  0.00007  Elements  Crust (%)  Soil (%)  Surface sediments (%)  Surface water (ppm)  O  46–50  49      Si  26–27  33      Al  7.5–8.3  7.1  6.72  0.72084  Fe  4.7–5.8  4  2.61  0.80041  Ca  3.39–5.2  1.5  1.2    K  2.3–2.64  1.4  2.46    Na  1.7–2.4  0.15  2.39    Mg  1.87–2.8  0.5  1.24  6.158  Ti  0.45–0.64  0.5  0.3186    Cl  0.13–0.19  0.01      P  0.09–0.12  0.08  0.05485    C  0.02–0.09  2      Mn  0.08–0.10  0.1  0.06239  0.01697  S  0.026–0.048  0.07      N  0.002–0.003  0.2      Cr  0.01–0.03  0.007  0.00997  0.00588  F  0.054–0.059  0.02      Ni  0.008–0.009  0.005  0.00829  0.00188  V  0.009–0.019  0.009  0.00547    Co  0.0018–0.0025  0.0008  0.00114  0.00045  Cu  0.005–0.006  0.003  0.0108  0.00626  Zn  0.007–0.009  0.0009  0.0388  0.010943  Pb  0.0012–0.0016  0.0035  0.00547  0.00807  As  0.00018–0.00022  0.0006  0.000885  0.003108  Br  0.00021–0.00025  0.001      Cd  0.000015–0.00002  3.5E-05  0.0000778  0.00007  Sources: Gaillardet, Viers and Dupré 2003; Yang et al.2003; JeffersonLab 2007; Liu et al.2007; Landaud, Helinck and Bonnarme 2008; Viers, Dupré and Gaillardet 2009; Feng et al.2010; Lin et al.2012; Song et al.2013; Gao et al.2016; Song et al.2017. View Large Table 3. Major clay minerals composition and content (%) in river sediments. Minerals  Pearl  Pearl River  Huanghe  Changjiang  Changjiang  Molecular    River  estuary  River  River  estuary  formula  Kaolinite  46  40  10  16  10  Al2Si2O5(OH)4              Clinochlore: (Mg5Al)(AlSi3)O10(OH)8              Chamosite: (Fe5Al)(AlSi3)O10(OH)8  Chlorite  25  28  16  12  26                Nimite: (Ni5Al)(AlSi3)O10(OH)8              Pennantite: (Mn,Al)6(Si,Al)4O10(OH)8  Illite  26  26  62  66  58  (K,H3O)(Al,Mg,Fe)2(Si,Al)4O10[(OH)2,(H2O)              Montmorillonite: (Na,Ca)0.33(Al,Mg)2(Si4O10)(OH)2·nH2O  Smectite  <2  <6  <12  <6  3                Nontronite: Na0.3Fe2((Si,Al)4O10)(OH)2·nH2O              Saponite: Ca0.25(Mg,Fe)3((Si,Al)4O10)(OH)2·nH2O  Minor              mineral  <1  <1  NP  NP  <3  FeS2  (Pyrite)              Minerals  Pearl  Pearl River  Huanghe  Changjiang  Changjiang  Molecular    River  estuary  River  River  estuary  formula  Kaolinite  46  40  10  16  10  Al2Si2O5(OH)4              Clinochlore: (Mg5Al)(AlSi3)O10(OH)8              Chamosite: (Fe5Al)(AlSi3)O10(OH)8  Chlorite  25  28  16  12  26                Nimite: (Ni5Al)(AlSi3)O10(OH)8              Pennantite: (Mn,Al)6(Si,Al)4O10(OH)8  Illite  26  26  62  66  58  (K,H3O)(Al,Mg,Fe)2(Si,Al)4O10[(OH)2,(H2O)              Montmorillonite: (Na,Ca)0.33(Al,Mg)2(Si4O10)(OH)2·nH2O  Smectite  <2  <6  <12  <6  3                Nontronite: Na0.3Fe2((Si,Al)4O10)(OH)2·nH2O              Saponite: Ca0.25(Mg,Fe)3((Si,Al)4O10)(OH)2·nH2O  Minor              mineral  <1  <1  NP  NP  <3  FeS2  (Pyrite)              NP, not provided. Sources: Liu et al.2007; Wang et al.2006; Yang et al.2003. View Large BIO-GEOCHEMICAL TRANSFERS OF THE KEY ELEMENTS INVOLVED IN WATER BLACKENING AND ODORIZAITON Black color formation via metal precipitates Black matter in urban rivers comprises black metallic precipitates and precipitates of brown, green, or other colors that together form a dark color. In O2-depleted surface waters, metals precipitate with sulfide and stain the water black (Fig. 1; Table 4; Nealson and Little 1997; Metzger et al.2014). Common metals such as iron and nickel form black or dark sulfides. Iron, nickel and copper sulfides are the thermodynamically most favorable precipitates (Table 4). Stahl (1979) investigated a black water lake in Illinois and demonstrated that ferrous sulfide was responsible for the black color. Mixed minerals such as copper–iron sulfides have been observed as well, for example, in the Danube River Basin (Brankov, Milijašević and Milanović 2012). Copper and other heavy metal precipitates were also detected in the Pearl River Estuary, China (Fang, Li and Zhang 2005) as well as the Reno River watershed, Italy (Ferronato et al.2013). In reduced environments, Mn exists as soluble Mn2+, due to its low affinity for sulfur, and dose not precipitate (Nealson and Little 1997). Figure 1. View largeDownload slide The microbially mediated reduction of Fe/Mn-minerals in urban rivers. Figure 1. View largeDownload slide The microbially mediated reduction of Fe/Mn-minerals in urban rivers. Table 4. Thermodynamics of some black or dark metal mineral reactions. Net reaction    ΔG°΄/M  ΔG°΄/S      (kJ mol−1)  (kJ mol−1)  Reductive environments      SO42− + H3C–COO− + 3 H+  →  HS− + 2 HCO3− + 3 H+  n/a  −48  Fe2+ + HS− + H+  →  FeSa + 2 H+  −231  −231  8 FeOOHb + 9 H3C–COO− + 8 SO42− + 25 H+  →  8 FeS + 18 HCO3− + 18 H+ + 12 H2O  −93  −93  2 FeOOH + 3 HS− + 3 H+  →  FeS + FeS2c + 4 H2O  −74  −50   Fe3O4 + 4 HS− + 4 H+  →  2 FeS + FeS2 + 4 H2O  −61  −46  FeS + S0  →  FeS2  −60  −30  4 FeOOH + 6 HS− + 6 H+  →  FeS2 + Fe3S4d + 8 H2O  −57  −38  2 FeOOH + 3 HS− + 3 H+  →  2 FeS + S0 + 4 H2O  −45  −30  Fe3O4e + 4 HS− + 4 H+  →  Fe3S4 + 4 H2O  −38  −28  9 FeS + 5 HS− + 5 H+  →  3 FeS2 + 2 Fe3S4  −2  −1  4 S0 + H3C–COO− + H+ + 4 H2O  →  4 HS− + 2 HCO3− + 6 H+  n/a  −2  24 FeOOHa + H3C–COO− + H+  →  8 Fe3O4 + 2 HCO3− + 2 H+ + 12 H2O  −5  n/a  8 FeOOH + H3C–COO− + 17 H+  →  8 Fe2+ + 2 HCO3− + 2 H+ + 12 H2O  186  n/a  Ni2+ + HS− + H+  →  NiSf + 2 H+  −184  −184  Cu2+ + HS− + H+  →  CuSg + 2 H+  −171  −171  CuS + S0  →  CuS2h  −33  −16  Pb2+ + HS− + H+  →  PbSi + 2 H+  −126  −126  Zn2+ + HS− + H+  →  ZnSj + 2 H+  −106  −106  Mn2+ + HS− + H+  →  MnSk + 2 H+  −42  −42  MnS + S0  →  MnS2l  167  167  Oxidative environments  ΔG°΄/Fe  ΔG°΄/ox    (kJ mol−1)  (kJ mol−1)  2 Cr2O3m + 3 O2 + 4 H2O  →  4 HCrO4− + 4 H+  −1094  −1459  10 FeS + 6 NO3− + 6 H+ + 2 H2O  →  10 FeOOH + 10 S0 + 3 N2  −250  −417  4 FeS + 3 O2 + 5 H2O  →  4 FeOOH + 4 S0 + 3 H2O  −270  −359  HS− + MnO2n + 3 H+  →  S0 + Mn2+ + 2 H2O  n/a  −130  2 FeS + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ 2 S0 + 2 H2O  −150  −100  2 FeS2 + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ + 4 S0 + 2 H2O  −90  −60  Net reaction    ΔG°΄/M  ΔG°΄/S      (kJ mol−1)  (kJ mol−1)  Reductive environments      SO42− + H3C–COO− + 3 H+  →  HS− + 2 HCO3− + 3 H+  n/a  −48  Fe2+ + HS− + H+  →  FeSa + 2 H+  −231  −231  8 FeOOHb + 9 H3C–COO− + 8 SO42− + 25 H+  →  8 FeS + 18 HCO3− + 18 H+ + 12 H2O  −93  −93  2 FeOOH + 3 HS− + 3 H+  →  FeS + FeS2c + 4 H2O  −74  −50   Fe3O4 + 4 HS− + 4 H+  →  2 FeS + FeS2 + 4 H2O  −61  −46  FeS + S0  →  FeS2  −60  −30  4 FeOOH + 6 HS− + 6 H+  →  FeS2 + Fe3S4d + 8 H2O  −57  −38  2 FeOOH + 3 HS− + 3 H+  →  2 FeS + S0 + 4 H2O  −45  −30  Fe3O4e + 4 HS− + 4 H+  →  Fe3S4 + 4 H2O  −38  −28  9 FeS + 5 HS− + 5 H+  →  3 FeS2 + 2 Fe3S4  −2  −1  4 S0 + H3C–COO− + H+ + 4 H2O  →  4 HS− + 2 HCO3− + 6 H+  n/a  −2  24 FeOOHa + H3C–COO− + H+  →  8 Fe3O4 + 2 HCO3− + 2 H+ + 12 H2O  −5  n/a  8 FeOOH + H3C–COO− + 17 H+  →  8 Fe2+ + 2 HCO3− + 2 H+ + 12 H2O  186  n/a  Ni2+ + HS− + H+  →  NiSf + 2 H+  −184  −184  Cu2+ + HS− + H+  →  CuSg + 2 H+  −171  −171  CuS + S0  →  CuS2h  −33  −16  Pb2+ + HS− + H+  →  PbSi + 2 H+  −126  −126  Zn2+ + HS− + H+  →  ZnSj + 2 H+  −106  −106  Mn2+ + HS− + H+  →  MnSk + 2 H+  −42  −42  MnS + S0  →  MnS2l  167  167  Oxidative environments  ΔG°΄/Fe  ΔG°΄/ox    (kJ mol−1)  (kJ mol−1)  2 Cr2O3m + 3 O2 + 4 H2O  →  4 HCrO4− + 4 H+  −1094  −1459  10 FeS + 6 NO3− + 6 H+ + 2 H2O  →  10 FeOOH + 10 S0 + 3 N2  −250  −417  4 FeS + 3 O2 + 5 H2O  →  4 FeOOH + 4 S0 + 3 H2O  −270  −359  HS− + MnO2n + 3 H+  →  S0 + Mn2+ + 2 H2O  n/a  −130  2 FeS + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ 2 S0 + 2 H2O  −150  −100  2 FeS2 + 3 MnO2 + 6 H+  →  2 FeOOH + 3 Mn2+ + 4 S0 + 2 H2O  −90  −60  Black or dark minerals: airon sulfide, bgoethite, cpyrite, dgreigite, emagnetite, fmillerite, gcovellite, hα-chalkosite, ilead sulfide, jsphalerite (disulfide not known for Zn and Pb), kalabandite (pink, orange or green), lhauerite, meskolaite, npyrolusite (light grey). M, metal; n/a, not applicable; ox. oxidant. View Large While iron sulfide formation is spontaneous, microorganisms such as Geobacter, Geothrix, Rhodoferax and Shewanella can harvest the released energy for their cell growth (Fig. 1; Lovley 1991). Thermodynamically, formation of FeS is favored followed by FeS2 and Fe3S4 (Table 4). Greigite (Fe3S4) is formed in excess of sulfide. Elemental sulfur (S0) and polysulfide (Sn2−) formation are thought to be the intermediate steps leading to pyrite and greigite (Table 4; Rickard 1975; Luther 1991). Similar to bio-geochemical transfers in marine sediments, the FeS and FeS2 as well as other metals play central roles in the sulfur cycle in urban rivers (Schippers and Jørgensen 2002). That is, metals and especially abundant iron, are reduced by organic pollutants, such as volatile fatty acids (such as acetic, butyric and propionic acids), or other reducing equivalents (Table 4). If iron then again enters oxidizing zones, for example by currents or shift of oxic zones, it can be re-oxidized by dissolved oxygen, nitrate, or manganese oxides. Iron oxides, such as goethite, can also be reduced biologically. Humic substances enhance the bioavailability of insoluble Fe(III) oxides as electron acceptors and therefore improve the thermodynamics of biological iron reduction (Lovley et al.1996, 1998). Quinone moieties in humic substances serve as electron shuttles in Fe(III)-respiring microorganisms, e.g. Ferribacterium limneticum and Geobacter metallireducens, accelerating the rate of both Fe(III) oxide reduction in river sediments and contaminant oxidation coupled to Fe(III) reduction (Lovley et al.1996; Finneran and Lovley 2001; Nevin and Lovley 2000). Other iron reducers such as Shewanella oneidensis (Venkateswaran et al.1999), Paludibaculum fermentans (Kulichevskaya et al.2014) and Anaeromyxobacter dehalogenans (Sanford, Cole and Tiedje 2002) are able to utilize a number of different electron donors including sugars and long chain fatty acids. The broad variety of electron acceptors and donors used by iron reducers makes these microorganisms ubiquitous in freshwater sediments. In sediments, pyrite (FeS2) is oxidized abiotically at mineral interfaces, for example between FeS2 and MnO2 (Table 4; Schippers and Jørgensen 2002). Immediate products of this oxidation are thiosulfate and polythionates, which can be further oxidized to sulfate by manganese-reducing bacteria (Jørgensen and Nelson 2004). Additionally, Fe(II) and Mn(II) were released from sediment pore waters to form black oxides in anoxic–oxic water interface zone of the black rivers (Atkinson et al.2007). Small amounts of dissolved ferrous iron released from FeS2 in sediments can be re-oxidized by oxygen, nitrate and MnO2 and precipitated as black magnetite in river beds. After reentry in oxic zones or through mediation, reduced Mn(II) is recycled to MnO2 via microbial oxidation in the presence of, even trace amounts of, O2 or nitrate in surface waters (Boogerd and de Vrind 1987; Marcus et al.2017). Nitrate has therefore been suggested as a cost-effective remediation method for black urban rivers (He et al.2017). Odorous volatile compounds Sulfur compounds Odorous compounds in urban rivers are volatile organic and inorganic compounds. Volatile sulfur compounds generated from microbial sulfate reduction or degradation of sulfur-containing organic matter normally have an unpleasant odor, including that of inorganic H2S and organic sulfides (Kadota and Ishida 2003). Sources of such reduced sulfur compounds vary. River deltas discharging into oceans often experience seawater influx due to tidal activity. Therefore, sulfide, as a result of microbial sulfate reduction, is detected in significant amounts in urban rivers connecting to major deltas in coastal areas, particularly in the Pearl River estuary where 3 mM sulfide in the sediment was reported as a result of sulfate reduction (Fang et al.2005). Most sulfide in estuaries is produced by sulfate reducing microorganisms using dissimilatory sulfate reduction for respiration. Sulfate reducing microorganisms are broadly dispersed across the prokaryotic phylogenetic tree but are often found among δ-Proteobacteria and Firmicutes as well as Archaea (Zhou et al.2011). In addition to the production of odorous H2S, sulfate reduction may eventually lead to the formation of black iron sulfide species even with trace amounts of iron in the water and sediment (Wu et al.2016). Another source of sulfide in urban rivers is organic sulfur (Giordano et al.2005). It enters the sulfur cycle via assimilatory sulfate reduction (Fig. 2). The amount of sulfur in domestic waste streams or other anthropogenic sources, however, is negligible compared with the massive cyanobacterial blooms in the surface waters (Zhang et al.2010). CO2-fixing cyanobacteria frequently are the main source of organic matter in surface waters. They assimilate sulfur via cysteine/methionine biosynthetic pathways (Fig. 2). This organic sulfur is subsequently released in the form of volatile organic sulfur compounds, which include thiols and thioethers in methylated sulfide species, e.g. methyl sulfide, dimethyl sulfide and dimethyl disulfide, as byproducts—all of which are characterized by their notoriously bad smell (Bentley and Chasteen 2004). Figure 2. View largeDownload slide Methionine biosynthesis (top) and degradation (bottom) proceed via cysteine. The biosynthetic pathway requires ATP for sulfate reduction to sulfide, producing the intermediate 3΄-phosphoadenosine-5΄-phosphosulfate (PAPS). Together with serine, sulfide then forms cysteine and ultimately methionine. Under anaerobic conditions, methionine is degraded to taurine, CO2 and sulfide in exchange for sulfite. Figure 2. View largeDownload slide Methionine biosynthesis (top) and degradation (bottom) proceed via cysteine. The biosynthetic pathway requires ATP for sulfate reduction to sulfide, producing the intermediate 3΄-phosphoadenosine-5΄-phosphosulfate (PAPS). Together with serine, sulfide then forms cysteine and ultimately methionine. Under anaerobic conditions, methionine is degraded to taurine, CO2 and sulfide in exchange for sulfite. Nitrogen and organic carbon compounds Nitrogen compounds are the largest group of malodorous compounds generated by proteolytic microorganisms, and their smells range from that of ammonia to the typical smell of corpse decomposition (Wang et al.2017). Microbiogenic malodorous nitrogen compounds include organic amines such as cadaverine (1,5-pentanediamine) and putrescine (1,4-butanediamine). Cadaverine is produced via decarboxylation of lysine, whereas putrescine is a product of degradation of ornithine, an essential building block of bacterial cell walls (Wunderlichová et al.2014; Ma et al.2017). Other abundant malodorous nitrogen compounds are volatile alkylated amines of characteristic fishy smell, which is sensed even in trace amounts. These compounds comprise alkylated amines, such as methylamine, dimethylamine, trimethylamine, ethaneamine, propaneamine and butaneamine (Ge et al.2011). Methylamines are degradation products of N-methylated amino acids, with glycine, betaine, choline, trimethylamine and carnitine as their natural precursors in biomass, which, in turn, is introduced into urban rivers via wastewater or algal blooms (Ikawa and Taylor 1973). All organisms are able to hydrolyze proteins using proteases as this is an essential part of their metabolism. Cell internal proteolysis is necessary in every organism, for example to tune its enzymatic machinery to novel environmental conditions or to control vital cell functions. Microorganisms specifically feeding on peptides can be isolated using casamino acids and trypticase peptone media, and often yield strains closely related to Clostridium species when grown anaerobically. Such proteolytic microorganisms are ubiquitous in anaerobic and aerobic environments alike. Examples of anaerobic environments are rumens (Blackburn and Hobson 1962), anaerobic digesters (Abendroth et al.2015), peat bogs (Juottonen et al.2005) and rice paddies (Weber et al.2001). Typical aerobic environments are many processed food products such as cabbage (Borla, Davidovich and Roura 2010) and dairy products (Frazier and Rupp 1931). Despite the presence of proteolytic microorganisms, none of the mentioned environments are known for their obnoxious smell. The reason is that protein concentrations are either relatively low or, in the case of food products, aerobic conditions prevail. When oxygen is absent, alkylated amines cannot be further oxidized and serve as substrates for sulfate reducers or methanogens (Lovley and Klug 1983). However, since alkylated amines are gaseous or at least volatile, they often escape before slow growing anaerobic microorganisms are able to degrade them, causing the typical smell in surface waters. Odorous organic compounds without S and N elements are mostly VFAs, which are generated by anaerobic fermentation of organic pollutants in the waste streams or of decomposed compounds produced by algal blooms (Verstraete et al.1996; Pham et al.2012). VFAs in urban rivers play critical roles in coupling organic carbon compounds with iron and sulfur cycles in the surface water (Fig. 4). For example, when sulfate is present, VFAs can be further used as electron donors by sulfate reducing microorganisms and produce malodorous sulfide (Hao et al.2009). Figure 3. View largeDownload slide Principal coordinate analysis (PCoA with Bray–Curtis distance matrices) of six surface-water sediment samples collected from three different geographic areas, i.e. Pearl River Delta, Yangtze River Delta and Dongting area. This figure is plotted with published 16S rRNA gene-sequencing data (Wang et al2012; Liu et al2014; He et al2017; Huang et al2017). The trophic states of the six surface waters are: eutrophic for Taihu Lake and Chebei River (black-odorous river); mesotrophic for Dongting Lake, Huangda Lake and Pearl River; and oligotrophic for Liuxi Reservoir. Figure 3. View largeDownload slide Principal coordinate analysis (PCoA with Bray–Curtis distance matrices) of six surface-water sediment samples collected from three different geographic areas, i.e. Pearl River Delta, Yangtze River Delta and Dongting area. This figure is plotted with published 16S rRNA gene-sequencing data (Wang et al2012; Liu et al2014; He et al2017; Huang et al2017). The trophic states of the six surface waters are: eutrophic for Taihu Lake and Chebei River (black-odorous river); mesotrophic for Dongting Lake, Huangda Lake and Pearl River; and oligotrophic for Liuxi Reservoir. MICROBIAL ECOLOGY OF PRISTINE AND POLLUTED FRESHWATER ENVIRONMENTS Urban river pollution affects microbial communities in water and sediments with measurable effects in the short (Schöll and Szövényi 2011) and long term (Ibekwe, Ma and Murinda 2016; Lu, Chen and Zheng 2017). This makes microbial community analysis an additional monitoring tool for water quality (Drury, Rosi-Marshall and Kelly 2013; García-Armisen et al.2014; Li, Sharp and Drewes 2016; Xie et al.2016; Köchling et al.2017). Diversity (Drury, Rosi-Marshall and Kelly 2013; Staley et al.2013), richness (Lin et al.2014) and variability (Lu, Chen and Zheng 2017) of microbial communities have been affected by anthropogenic pollutants. As expected, coliform growth is correlated with fecal anthropogenic contamination, for example in large rivers such as the Danube (Hoch et al.1996; Kirschner et al.2009) and the Mississippi (Staley et al.2013), as well as smaller rivers such as the Jaboatão River in Brazil (Köchling et al.2017), the Reno River in Italy (Ferronato et al.2013) and small creeks of a California watershed (Ibekwe, Ma and Murinda 2016). A clear impact of treated wastewater on community composition and metabolism was reported for the Taif River in Saudi Arabia, where pristine samples showed a higher representation of carbohydrate metabolic genes along with fatty acid and amino acid anabolic genes as opposed to samples impacted by wastewater (Li, Sharp and Drewes 2016). The latter were enriched in genes associated with nitrogen and sulfur metabolism, as would be expected in nutrient rich environments. Inverse metabolic patterns were reported for river sediments in China where energy-, carbohydrate- and amino acid-related genes were enriched or equal to pristine control sediments (Lu, Chen and Zheng 2017). Algal growth is often considered to be linked to anthropogenic contamination in freshwater systems such as Taihu Lake in China (Huang et al.2017) or the Zenne River in Belgium (García-Armisen et al.2014). However, the mechanisms by which pollution and algal growth are connected are not always clear. For example, Huang et al. (2017) found that phosphate as well as organic matter concentration were correlated with cyanobacterial growth in the Taihu Lake, whereas in the Danube River, Kirschner et al. (2009) identified only a link to organic matter but not any other of the factors investigated, such as phosphate, nitrogen and temperature. This suggests that cyanobacteria live heterotrophically or that low bioavailability of inorganic nutrients are responsible for algal blooms. Despite the ongoing DNA sequencing revolution, it is not clear which factors shape river sediment communities. Some recent attempts indicate that, indeed, organic matter released into urban rivers by sewage streams promotes growth of certain microbial lineages such as Acinetobacter, Flaviobacterium, Thauera and Rhodococcus in the Zenne River flowing through the Brussels metropolitan area (García-Armisen et al.2014). A similar correlation between organic matter and Cyanobacteria was linked to fecal coliforms and Enterococci in the Danube (Kirschner et al.2009). The dependence of fecal coliforms on environmental factors, however, was stronger in the water column than in the sediments studied in selected creeks of a southern California watershed (Ibekwe, Ma and Murinda 2016). In the Rhône River prodelta, microbial variation could be explained by organic matter as well (Fagervold, et al.2014). In addition to organic matter, Ji et al. (2016) found that iron and sulfate concentrations as well as pH were associated with methanogenic networks identified in Amazonian lake sediments. The pioneering works of Zwart et al. (2002) and Newton et al. (2011) identified an appreciable bacterial diversity in freshwater systems showing that river and lake communities are similar. In addition, our own comparison of published 16S rRNA gene-sequencing data of three Chinese lakes, two rivers and one reservoir from three distant areas shows that geographical location best explained the differences between the investigated freshwater environments (Fig. 3). Though microbial populations of contaminated and pristine sites of the same region cluster closely, communities still show some difference. Proteobacteria are the largest phylum in the prokaryotic tree of life and are therefore also highly abundant in river sediments—polluted or not. Nearly all river and lake sediments surveyed here harbored α-, β- and γ-Proteobacteria. The most prominent representative of freshwater α-Proteobacteria is the SAR11 clade (Pelagibacter; Salcher et al.2011; Savio et al.2015). Also Bacteroidetes were found in nearly all freshwater sediments. Together, these four groups cover 40% of all cultured prokaryotic species, making their dominance in freshwater sediments only natural. Consequently, microbial communities in freshwater sediments are often very similar at the phylum level (Ji et al.2016). Nonetheless, a study screening 68 publications of lake microbial communities using only high quality Sanger-sequencing data reported a large heterogeneity at lower taxonomic levels, termed tribes (Newton et al.2011). This finding was confirmed for Mississippi River sediments where only 12% of the identified operational taxonomic units (>97% sequence identity) were shared across all sites (Staley et al.2013). It is hence the less abundant groups, such as Acidobacteria, Actinobacteria, Verrucomicrobia, Chloroflexi, Planctomycetes, Gemmatimonadetes and Archaea, or tribes that may act as distinctive indicators for metabolic processes. Many such tribes are uncultured representatives of freshwater environments and do not match with Linnaean taxonomic boundaries. While broad surveys of our drinking water resources need to be continued, it remains unclear how under-represented parts of microbial communities adapt to pollution. Indeed, a recent metagenomic survey of a wastewater impacted river showed that small community factions are major hubs in microbial assemblages (Li, Sharp and Drewes 2016). This is an observation that has also been made in the pristine Lake Cadagno, Switzerland, where 0.3% of the cells in the lake were responsible for 40% of the substrate turnover (Musat et al.2008). In conclusion, more studies on natural environments are necessary to understand the above mentioned discrepancies and to establish a baseline for future research on pollution affected environments. A SCENARIO OF THE BIO-GEOCHEMICAL PROCESS IN BLACKENING AND ODORIZAITON OF URBAN RIVERS In the scenario depicted in Fig. 4, organic matter originates either directly from anthropogenic sources, e.g. waste streams and other non-point source pollution, or from decomposition of cyanobacteria biomass. These are the main suspects in generating odors by production of volatile (organic) sulfides, odorous amines and VFAs (Van Neste et al.1987; Ginzburg et al.1998; Hu et al.2007; Zhang et al.2010). Organic matter is also the major source of reducing equivalents for the reduction of sulfate and iron to produce black minerals such as iron sulfides, which link the sulfur and iron cycles in urban rivers (Fig. 4; Berner et al.1985; Lovley 1987). Therefore, input of organic matter into urban rivers is likely the key factor for triggering water blackening and odorization. Sulfur input from sediment, seawater, or decomposing algal biomass is directly involved in the formation of black- and odorous-matter in urban rivers, e.g. via formation of sulfide species such as H2S and iron sulfides. Alkylated sulfides are often byproducts of cyanobacterial metabolism and biomass degradation and their volatility makes them strong odorous constituents of some urban rivers’ stench. Sulfate reduction by diverse sulfate reducing microorganisms dominates in sediments because their redox potential confers growth advantages to these microorganisms over their competitors. For example, the standard redox potential of sulfate reduction (−217 mV) is slightly more positive than that of hydrogenotropic methanogenic process (−240 mV). This and the higher energy gain compared with iron reduction suggest that sulfur is the link between the different element cycles as shown for Black Sea sediments (Siegert et al.2013). Figure 4. View largeDownload slide Scenarios describing bio-geochemical transfers of major black and odorous elements/compounds in urban rivers. In the process of organic matter degradation, odors result from production of volatile organic sulfur compounds, methylamines and VFAs, which escape into the atmosphere. In addition, organic matter supplies reducing power for SO42− and Fe(III) to form FeSx. The processes for black/odorous compounds are shown with solid lines, and others with dotted lines. DMS, dimethyl sulfide; VFAs, volatile organic sulfur compounds; VOSCs, volatile organic sulfur compounds. Figure 4. View largeDownload slide Scenarios describing bio-geochemical transfers of major black and odorous elements/compounds in urban rivers. In the process of organic matter degradation, odors result from production of volatile organic sulfur compounds, methylamines and VFAs, which escape into the atmosphere. In addition, organic matter supplies reducing power for SO42− and Fe(III) to form FeSx. The processes for black/odorous compounds are shown with solid lines, and others with dotted lines. DMS, dimethyl sulfide; VFAs, volatile organic sulfur compounds; VOSCs, volatile organic sulfur compounds. CONCLUSIONS AND FUTURE PERSPECTIVES The blackening and odorization of urban rivers is a complex bio-geochemical process involving five key elements, i.e. Fe, Mn, S, N and C. Outstanding issues include the following. (i) While we propose several mechanisms that contribute to blackening and odorization of urban rivers such as organic matter degradation and metal precipitation, there is no evidence yet that these are indeed the driving factors. In the past, measures to counter river pollution were taken, such as widespread treatment of industrial wastewater in the Pearl River Delta and sediment removal, yet they only mitigated the problem for a short time. Evidence for our hypothesized mechanisms needs to be collected in order to take targeted action. Gathering this evidence requires application of standard tests to assess water quality (listed in Table 5) along with novel molecular techniques and may involve the development of new methods that are more efficient. As shown in this review, the suspected blackening elements (metal sulfides) and three odor-forming elements (S, N and C) should be first targets for water quality analysis investigating blackening and odor formation in the urban rivers. Table 5. Routine methods for surface water quality assessment. Analyte  Method  References  Dissolved inorganic carbon  Spectrophotometry/potentiometry/conductimetry  Oshima et al.2001; Carlson 1978; Linares et al.1989  Dissolved organic carbon  Chemical oxidation  Sharp 1973    Ultraviolet oxidation  Beattie et al.1961; Armstrong et al.1966    High-temperature combustion  Sharp 1973; Salonen 1979  Chemical oxygen demand  Dichromate oxidation method  Moore et al.1949; Jirka and Carter 1975    Potassium permanganate oxidation method  Korenaga 1980  Metal content (Fe, Mn, Cu, Zn, etc.)  Spectroscopic analysis method (inductively coupled plasma mass spectrometry)  Houk et al. 1989  Total dissolved nitrogen  Alkaline persulfate digestion  Solorzano and Sharp 1980    High temperature oxidation  Suzuki et al.1985  Dissolved inorganic nitrogen  Phenol hypochlorite reaction method (NH3/NH4+)  Bolleter et al.1961    Nessler's reagent spectrophotometry (NH3/NH4+)  Vanselow 1940; Leonard 1963    Ion chromatography (NO2−, NO3−)  Gjerde et al.1979    Colorimetry (NO2−)  APHA 1998    Ultraviolet spectrophotometry (NO3−)  Hoather and Rackham 1959  Dissolved organic nitrogen  High-temperature catalytic oxidation  Badr et al.2003  Dissolved and precipitated sulfides  CuS colloidal solution method  Cord-Ruwisch 1985  Dissolved sulfate  Turbidimetry  Tabatabai 1974  Volatile organic sulfur compounds  Chromatography analysis method (gas chromatography–sulfur chemiluminescence detection)  Sun et al.2014    Chromatography analysis method (gas chromatography–mass spectrometry)  Van Langenhove et al.1985  Element content (C, H, N, S)  Elemental analysis  Kirsten 1971; Fadeeva et al.2008  Analyte  Method  References  Dissolved inorganic carbon  Spectrophotometry/potentiometry/conductimetry  Oshima et al.2001; Carlson 1978; Linares et al.1989  Dissolved organic carbon  Chemical oxidation  Sharp 1973    Ultraviolet oxidation  Beattie et al.1961; Armstrong et al.1966    High-temperature combustion  Sharp 1973; Salonen 1979  Chemical oxygen demand  Dichromate oxidation method  Moore et al.1949; Jirka and Carter 1975    Potassium permanganate oxidation method  Korenaga 1980  Metal content (Fe, Mn, Cu, Zn, etc.)  Spectroscopic analysis method (inductively coupled plasma mass spectrometry)  Houk et al. 1989  Total dissolved nitrogen  Alkaline persulfate digestion  Solorzano and Sharp 1980    High temperature oxidation  Suzuki et al.1985  Dissolved inorganic nitrogen  Phenol hypochlorite reaction method (NH3/NH4+)  Bolleter et al.1961    Nessler's reagent spectrophotometry (NH3/NH4+)  Vanselow 1940; Leonard 1963    Ion chromatography (NO2−, NO3−)  Gjerde et al.1979    Colorimetry (NO2−)  APHA 1998    Ultraviolet spectrophotometry (NO3−)  Hoather and Rackham 1959  Dissolved organic nitrogen  High-temperature catalytic oxidation  Badr et al.2003  Dissolved and precipitated sulfides  CuS colloidal solution method  Cord-Ruwisch 1985  Dissolved sulfate  Turbidimetry  Tabatabai 1974  Volatile organic sulfur compounds  Chromatography analysis method (gas chromatography–sulfur chemiluminescence detection)  Sun et al.2014    Chromatography analysis method (gas chromatography–mass spectrometry)  Van Langenhove et al.1985  Element content (C, H, N, S)  Elemental analysis  Kirsten 1971; Fadeeva et al.2008  View Large (ii) What, if any, are the core microbial communities, taxonomically and physiologically, mediating bio-geochemical transfers of the key elements in black-odorous urban rivers? Despite the progress in studies of metabolism and element cycles in surface waters, many puzzles remain, for example, the discrepancy in some reports and our own investigations showing that pollution sometimes does not affect microbial communities. There is a need for better understanding of the core communities coupling all these element cycles in black-odorous urban rivers. Current meta-omic technologies may help to provide in-depth insights. (iii) What is the role of minor elements, e.g. Cu and Zn, in the blackening and odorization of urban rivers? Hitherto, very few studies have investigated their contribution to blackening and odorization in surface waters. Whether these trace elements play critical roles in connecting the Fe, Mn, S, N and C cycles warrants future investigation. (iv) New water quality standards to address blackening and odorous surface waters need to be developed in China. 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FEMS Microbiology EcologyOxford University Press

Published: Mar 1, 2018

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