**Environmental Trace Elements**

**Chapter 6**

**Provisional chapter**

**Trace Elements in Suspended Particulate Matter and**

**Trace Elements in Suspended Particulate Matter and** 

DOI: 10.5772/intechopen.76471

**Sediments of the Cai River: Nha Trang Bay Estuarine**

**Sediments of the Cai River: Nha Trang Bay Estuarine** 

The distribution of particulate form of organic carbon (POC), Al, Fe, Ti, Li, Zn, Pb, U, Sc, Sn, Bi, Zr, Ba, As, Sr, W, V, Co, Cu, Ni, Mo, Cr, Mn, Ba, Sn, Sb, Hg, and Ag in the Cai river and Nha Trang Bay generally followed the distribution of total suspended matter (SPM) and was characterized by the most significant loss in the frontal zone of the estuary with highest horizontal gradients within the salinity interval of 8–20‰. The most part of these elements are supplied to the estuary with the Cai river discharge. Sedimentary Al, Fe, Ti, Li, Sc, Co, Cs, Zr, Cr, Zn, Co, Ni, Cu, Pb, Sn, and V are most likely controlled by the accumulation of their most fine-grained host minerals in sea floor depression of the bay. Sedimentary Bi, W, As, U, and Mo are mainly deposited with the coarse river material near the river mouth. The distribution of Ca, Sr, Mn, and Ba is largely controlled by the total inorganic carbon (TIC) content in the sediments. Metal form study revealed the highest percent contents of the labile forms for Mn, Co, and Pb in the sediments. The high levels of weak acid-soluble Pb and Co (30% and 43% of the total content in sediment, on average, respectively) contributes to a contamination problem in the Nha Trang Bay

**Keywords:** Vietnam, tropical estuary, trace elements, selective extraction, bioavailable

Southeast Asia has experienced a rapid and colossal economic growth with Vietnam being one of the fastest growing countries. Most development activities (e.g., industry, agriculture,

> © 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

© 2018 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use,

distribution, and reproduction in any medium, provided the original work is properly cited.

**System (South China Sea)**

**System (South China Sea)**

Sofia Koukina and Nikolay Lobus

Sofia Koukina and Nikolay Lobus

http://dx.doi.org/10.5772/intechopen.76471

**Abstract**

metal forms

**1. Introduction**

Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

which arises from the Cai River discharge.

#### **Trace Elements in Suspended Particulate Matter and Sediments of the Cai River: Nha Trang Bay Estuarine System (South China Sea) Trace Elements in Suspended Particulate Matter and Sediments of the Cai River: Nha Trang Bay Estuarine System (South China Sea)**

DOI: 10.5772/intechopen.76471

Sofia Koukina and Nikolay Lobus Sofia Koukina and Nikolay Lobus

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/intechopen.76471

#### **Abstract**

The distribution of particulate form of organic carbon (POC), Al, Fe, Ti, Li, Zn, Pb, U, Sc, Sn, Bi, Zr, Ba, As, Sr, W, V, Co, Cu, Ni, Mo, Cr, Mn, Ba, Sn, Sb, Hg, and Ag in the Cai river and Nha Trang Bay generally followed the distribution of total suspended matter (SPM) and was characterized by the most significant loss in the frontal zone of the estuary with highest horizontal gradients within the salinity interval of 8–20‰. The most part of these elements are supplied to the estuary with the Cai river discharge. Sedimentary Al, Fe, Ti, Li, Sc, Co, Cs, Zr, Cr, Zn, Co, Ni, Cu, Pb, Sn, and V are most likely controlled by the accumulation of their most fine-grained host minerals in sea floor depression of the bay. Sedimentary Bi, W, As, U, and Mo are mainly deposited with the coarse river material near the river mouth. The distribution of Ca, Sr, Mn, and Ba is largely controlled by the total inorganic carbon (TIC) content in the sediments. Metal form study revealed the highest percent contents of the labile forms for Mn, Co, and Pb in the sediments. The high levels of weak acid-soluble Pb and Co (30% and 43% of the total content in sediment, on average, respectively) contributes to a contamination problem in the Nha Trang Bay which arises from the Cai River discharge.

**Keywords:** Vietnam, tropical estuary, trace elements, selective extraction, bioavailable metal forms

## **1. Introduction**

Southeast Asia has experienced a rapid and colossal economic growth with Vietnam being one of the fastest growing countries. Most development activities (e.g., industry, agriculture,

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2018 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

human settlement, tourism, and transport) take place in coastal zones [1–4]. This hazard is increased by the high vulnerability of these areas to environmental changes. The Cai River and the Nha Trang Bay of the South China Sea are inhabited by unique biota. This region is now exposed to the multiple anthropogenic stressors such as human settlement, agriculture and aquaculture, tourism and transport [5].

highly stratified with the pronounced horizontal and vertical salinity gradients [5]. The climate seasonality and human activities (such as urbanisation, land use, damming, tourism, coastal construction, transportation, aquaculture and fisheries) expose the Nha Trang Bay to

Trace Elements in Suspended Particulate Matter and Sediments of the Cai River: Nha Trang Bay…

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115

The water and sediment samples were collected in the Cai River estuary and Nha Trang Bay in July 2013 along the salinity gradient at five locations for surface water layer (sts. 1, 3, 4, 7, 8, **Figure 1**) and at seven locations for surface sediments (sts. 2–8, **Figure 1**). The sampling stations were located in the riverine (st. 1), transitional (sts. 2–4) and marine (sts. 5–8) sub-zones

The surface water samples were obtained using a plastic Niskin bottle. The temperature, alkalinity and salinity of the water samples were measured on-board immediately after collection using portable conductivity apparatuses HI 98129 Combo and HI 98302 DIST 2 (Hanna Instruments, Germany). The suspended particulate matter was collected by filtering of water samples in an all-glass filtering system, on pre-weighted filters: 0.45 μm polycarbonate filters (Millipore-Isopore) for total suspended matter (TSM); combusted and pre-weighted glass fibre filters (Whatman GF/F) for particulate organic carbon (POC) and acid-clean cellulose filters (Millipore HA) for geochemical analyses. In the laboratory, all filters were rinsed with 250 ml Milli-Q water to remove salts and dried to constant weight

The surface sediment sampling, transportation and preparation procedures were performed

The surface sediment samples were subjected to grain size and mineral composition analyses.

The dissolved organic carbon in water samples was determined by high-temperature (at 680°С) thermocatalitic oxidation with dispersion-free IP detection. The total carbon (TC) contents in suspended particulate matter (SPM) and sediment samples were determined by dry burning at 900°С in oxygen flow and the total inorganic carbon (TIC) contents were

PO<sup>4</sup>

formed with the analyser TOC 5000-V-CPH (Shimudzu Co., Japan). The total organic carbon (TOC) contents were determined as a difference between TC and TIC contents in the

For the total Al, Fe, Ti, Ca, Na, Mn, Li, V, Cr, Co, Ni, Cu, Zn, As, Sr., Zr, Mo, Cd, Ag, Sn, Sb, Cs, Ba, Pb, Bi and U content analysis, the samples were subjected to the total acidic dissolution in

in an open system with further determination of element contents using

. The DOC, TC and TIC analyses were per-

using standard clean techniques that were described elsewhere [5].

The grain size analysis was performed by wet sieving [27, 28].

determined by dry burning at 200°С with H3

multiple pressures [14, 23–26].

**2.2. Field work**

of the dry season.

at 60°C.

**2.3. Analytical methods**

samples [10].

HNO<sup>3</sup> + HF + HClO<sup>4</sup>

Over the past two decades, comprehensive studies of the organic geochemistry patterns and contamination levels and trends in the Nha Trang Bay have been undertaken [6–14]. The most recent research of the abundance, distribution and speciation of the major and trace elements in the sediments allowed to track the fate of potential contaminants in the Cai River—Nha Trang Bay estuarine system along the salinity gradient [5, 15]. It was shown that most trace element contents were at natural levels and are derived from the composition of rocks and soils in the watershed. A severe enrichment of Ag was most likely derived from metal-rich detrital heavy minerals. Geochemical fractionation of the riverine material generally determined the metal enrichment in surface sediments along the salinity gradient. The parts of actually and potentially bioavailable forms were most elevated for Mn and Pb (up to 36 and 32% of total content, respectively). Overall, the most bioavailable parts of trace elements were associated with easily soluble amorphous Fe and Mn oxyhydroxides.

In estuarine region, suspended particulate matter (SPM) acts as a major carrier as trace elements get adsorbed on to major elements like Fe and Mn oxyhydroxides and organic matter and get precipitated, where coarse material may settle into the estuarine system as sediments and finer materials get transported into the ocean [16–18]. It is important to study major and trace elements, as excess input of these metals may settle into the estuary due to salinity gradient [19–22]. The present study summarises the data on the abundance, distribution, partition, speciation and bioavailability of major and trace elements in the suspended particulate matter (SPM) and surface sediments of the Cai River estuary under multiple stresses.

## **2. Materials and methods**

## **2.1. Environmental setting**

The Cai River, its estuary and the adjacent part of the Nha Trang Bay belong to the Central Southern Coastal Region of Vietnam (Khanh Hoa Province). The fresh river water (S < 0.1‰) and saline South China Sea water (S ≈ 36‰) form a major water-mixing zone. The fill dam built 8 km upstream from the river mouth limits the water exchange and marks the riverine boundary of the water-mixing zone (**Figure 1**). The Cai River estuary and Nha Trang Bay can be divided into three sub-zones: (1) river (S < 0.1‰), (2) transitional waters (estuary) (0.1‰ > S > 32‰) and (3) sea (bay) (S > 32‰). In the transitional waters, the salinity (S) increases from the river to the sea and from the surface to the bottom. The water column is highly stratified with the pronounced horizontal and vertical salinity gradients [5]. The climate seasonality and human activities (such as urbanisation, land use, damming, tourism, coastal construction, transportation, aquaculture and fisheries) expose the Nha Trang Bay to multiple pressures [14, 23–26].

### **2.2. Field work**

human settlement, tourism, and transport) take place in coastal zones [1–4]. This hazard is increased by the high vulnerability of these areas to environmental changes. The Cai River and the Nha Trang Bay of the South China Sea are inhabited by unique biota. This region is now exposed to the multiple anthropogenic stressors such as human settlement, agriculture

Over the past two decades, comprehensive studies of the organic geochemistry patterns and contamination levels and trends in the Nha Trang Bay have been undertaken [6–14]. The most recent research of the abundance, distribution and speciation of the major and trace elements in the sediments allowed to track the fate of potential contaminants in the Cai River—Nha Trang Bay estuarine system along the salinity gradient [5, 15]. It was shown that most trace element contents were at natural levels and are derived from the composition of rocks and soils in the watershed. A severe enrichment of Ag was most likely derived from metal-rich detrital heavy minerals. Geochemical fractionation of the riverine material generally determined the metal enrichment in surface sediments along the salinity gradient. The parts of actually and potentially bioavailable forms were most elevated for Mn and Pb (up to 36 and 32% of total content, respectively). Overall, the most bioavailable parts of trace elements were associated with easily soluble amorphous Fe and

In estuarine region, suspended particulate matter (SPM) acts as a major carrier as trace elements get adsorbed on to major elements like Fe and Mn oxyhydroxides and organic matter and get precipitated, where coarse material may settle into the estuarine system as sediments and finer materials get transported into the ocean [16–18]. It is important to study major and trace elements, as excess input of these metals may settle into the estuary due to salinity gradient [19–22]. The present study summarises the data on the abundance, distribution, partition, speciation and bioavailability of major and trace elements in the suspended particulate matter (SPM) and surface sediments of the Cai River estuary under

The Cai River, its estuary and the adjacent part of the Nha Trang Bay belong to the Central Southern Coastal Region of Vietnam (Khanh Hoa Province). The fresh river water (S < 0.1‰) and saline South China Sea water (S ≈ 36‰) form a major water-mixing zone. The fill dam built 8 km upstream from the river mouth limits the water exchange and marks the riverine boundary of the water-mixing zone (**Figure 1**). The Cai River estuary and Nha Trang Bay can be divided into three sub-zones: (1) river (S < 0.1‰), (2) transitional waters (estuary) (0.1‰ > S > 32‰) and (3) sea (bay) (S > 32‰). In the transitional waters, the salinity (S) increases from the river to the sea and from the surface to the bottom. The water column is

and aquaculture, tourism and transport [5].

114 Trace Elements - Human Health and Environment

Mn oxyhydroxides.

multiple stresses.

**2. Materials and methods**

**2.1. Environmental setting**

The water and sediment samples were collected in the Cai River estuary and Nha Trang Bay in July 2013 along the salinity gradient at five locations for surface water layer (sts. 1, 3, 4, 7, 8, **Figure 1**) and at seven locations for surface sediments (sts. 2–8, **Figure 1**). The sampling stations were located in the riverine (st. 1), transitional (sts. 2–4) and marine (sts. 5–8) sub-zones of the dry season.

The surface water samples were obtained using a plastic Niskin bottle. The temperature, alkalinity and salinity of the water samples were measured on-board immediately after collection using portable conductivity apparatuses HI 98129 Combo and HI 98302 DIST 2 (Hanna Instruments, Germany). The suspended particulate matter was collected by filtering of water samples in an all-glass filtering system, on pre-weighted filters: 0.45 μm polycarbonate filters (Millipore-Isopore) for total suspended matter (TSM); combusted and pre-weighted glass fibre filters (Whatman GF/F) for particulate organic carbon (POC) and acid-clean cellulose filters (Millipore HA) for geochemical analyses. In the laboratory, all filters were rinsed with 250 ml Milli-Q water to remove salts and dried to constant weight at 60°C.

The surface sediment sampling, transportation and preparation procedures were performed using standard clean techniques that were described elsewhere [5].

#### **2.3. Analytical methods**

The surface sediment samples were subjected to grain size and mineral composition analyses. The grain size analysis was performed by wet sieving [27, 28].

The dissolved organic carbon in water samples was determined by high-temperature (at 680°С) thermocatalitic oxidation with dispersion-free IP detection. The total carbon (TC) contents in suspended particulate matter (SPM) and sediment samples were determined by dry burning at 900°С in oxygen flow and the total inorganic carbon (TIC) contents were determined by dry burning at 200°С with H3 PO<sup>4</sup> . The DOC, TC and TIC analyses were performed with the analyser TOC 5000-V-CPH (Shimudzu Co., Japan). The total organic carbon (TOC) contents were determined as a difference between TC and TIC contents in the samples [10].

For the total Al, Fe, Ti, Ca, Na, Mn, Li, V, Cr, Co, Ni, Cu, Zn, As, Sr., Zr, Mo, Cd, Ag, Sn, Sb, Cs, Ba, Pb, Bi and U content analysis, the samples were subjected to the total acidic dissolution in HNO<sup>3</sup> + HF + HClO<sup>4</sup> in an open system with further determination of element contents using

To assess the chemical form of selected metals (Fe, Mn, Cr, Zn, Cu, Pb, Ni and Co) in the sediments, the samples were subjected to single chemical reagent (single-step) extraction procedures. The weak-acid-soluble (labile) metals were extracted using 25% acetic acid, the oxalate-soluble metals were extracted using ammonium oxalate-oxalic acid buffered solution at pH 3.2 (Tamm extraction), the pyrophosphate-soluble metals were extracted by 0.1 M sodium pyrophosphate. To isolate the weak acid-soluble metals, 15 ml of 25% acetic acid was added to 1.1 g of dry sample in polypropylene vials and shaken in a mechanical shaker for 6 h with acetic acid. Then, each extract with the sediment was filtrated into a 25 ml glass volumetric flask. The sediment on the filter was washed with 10 ml of distilled water and the wash water was added to the flask [27]. To isolate the amorphous iron oxides and their associated microelements, 50 ml of ammonium oxalate-oxalic acid buffered (Tamm) solution was added to 1.1 g of dry sample in the 250 ml flat bottom flask, shaken for 1 h and filtrated to a 250 ml glass volumetric flask. The sediment on the filter was washed with 10 ml of distilled water that was mixed with a small amount of oxalic acid and the wash water was added to the flask. Then, the filter with sediment was added to the sediment in the flat-bottom flask and subjected to one more repeated extraction and the extract was added to the 250 ml volumetric flask [30]. To isolate the organically bound metals, 15 ml of 0.1 M sodium pyrophosphate was added to 1.1 g of dry sample in polypropylene vials and shaken in a mechanical shaker for 15 min, left for 24 h and then filtrated to a 250 ml glass volumetric flask. The sediment on the filter was washed with 10 ml of 0.1 M sodium pyrophosphate and the wash water was added to the flask [31]. The metal contents in the extracts were further determined using an atomic absorption spectrometer (AAS) Hitachi 180–8 (Hitachi Co., Japan) in the Analytical Centre of Moscow State Lomonosov

Trace Elements in Suspended Particulate Matter and Sediments of the Cai River: Nha Trang Bay…

The relative accuracy of the analytical determinations was within the standard deviations that

The hydrology of the riverine sub-zone of the studied part of the Cai River—Nha Trang Bay estuarine system is strongly influenced by the fill dam. The surface fresh water layer flows seaward over the dam, while the upstream penetration of the near-bottom saline water lenses is blocked [5]. In July 2013, the salinity of the surface fresh-water layer varied from 0% to 36‰. The frontal zone of the contact of fresh and saline waters occurred downstream from the fill dam (sts. 2–3, **Figure 1**), where the horizontal salinity gradient was 3.5‰ per 1 km distance. The temperature (T) of the water column varied in narrow ranges and decreased from

, Stotal, K2

O5 , TiO2

O, CaO, MnO, Fe2

O3

http://dx.doi.org/10.5772/intechopen.76471

117

, Co, Ni, Cu, Sr., Sb, Cs, U)

, TOC

, Li, V, Cr, Zn,

were established by the certified reference materials (CRM) SDO-1 (Russia) (for SiO<sup>2</sup>

O, Al2 O3

**3.1. Abundance and distribution of major and trace elements in surface SPM**

University.

and TIC), SRM 521-84Р (Russia) (for Na<sup>2</sup>

and Mess-3 (Canada) (for Hg).

**3. Results and discussion**

As, Sr., Zr, Mo, Ag, Sn, Ba, Pb), AGV-2 (USA) (for MgO, P<sup>2</sup>

**Figure 1.** Location of study sites.

the ICP method on the Х-7 ICP-MS spectrometer (Thermo Scientific, USA) [29]. The detailed sample decomposition and analytical procedures are described elsewhere [16]. The Hg content was determined in the dry samples using a pyrolyse method on the RA-915+ spectrometer with background correction and a two-chamber atomiser PYRO-915+ (Lumex, Russia) [9, 14].

To assess the chemical form of selected metals (Fe, Mn, Cr, Zn, Cu, Pb, Ni and Co) in the sediments, the samples were subjected to single chemical reagent (single-step) extraction procedures. The weak-acid-soluble (labile) metals were extracted using 25% acetic acid, the oxalate-soluble metals were extracted using ammonium oxalate-oxalic acid buffered solution at pH 3.2 (Tamm extraction), the pyrophosphate-soluble metals were extracted by 0.1 M sodium pyrophosphate. To isolate the weak acid-soluble metals, 15 ml of 25% acetic acid was added to 1.1 g of dry sample in polypropylene vials and shaken in a mechanical shaker for 6 h with acetic acid. Then, each extract with the sediment was filtrated into a 25 ml glass volumetric flask. The sediment on the filter was washed with 10 ml of distilled water and the wash water was added to the flask [27]. To isolate the amorphous iron oxides and their associated microelements, 50 ml of ammonium oxalate-oxalic acid buffered (Tamm) solution was added to 1.1 g of dry sample in the 250 ml flat bottom flask, shaken for 1 h and filtrated to a 250 ml glass volumetric flask. The sediment on the filter was washed with 10 ml of distilled water that was mixed with a small amount of oxalic acid and the wash water was added to the flask. Then, the filter with sediment was added to the sediment in the flat-bottom flask and subjected to one more repeated extraction and the extract was added to the 250 ml volumetric flask [30]. To isolate the organically bound metals, 15 ml of 0.1 M sodium pyrophosphate was added to 1.1 g of dry sample in polypropylene vials and shaken in a mechanical shaker for 15 min, left for 24 h and then filtrated to a 250 ml glass volumetric flask. The sediment on the filter was washed with 10 ml of 0.1 M sodium pyrophosphate and the wash water was added to the flask [31]. The metal contents in the extracts were further determined using an atomic absorption spectrometer (AAS) Hitachi 180–8 (Hitachi Co., Japan) in the Analytical Centre of Moscow State Lomonosov University.

The relative accuracy of the analytical determinations was within the standard deviations that were established by the certified reference materials (CRM) SDO-1 (Russia) (for SiO<sup>2</sup> , TOC and TIC), SRM 521-84Р (Russia) (for Na<sup>2</sup> O, Al2 O3 , Stotal, K2 O, CaO, MnO, Fe2 O3 , Li, V, Cr, Zn, As, Sr., Zr, Mo, Ag, Sn, Ba, Pb), AGV-2 (USA) (for MgO, P<sup>2</sup> O5 , TiO2 , Co, Ni, Cu, Sr., Sb, Cs, U) and Mess-3 (Canada) (for Hg).

## **3. Results and discussion**

the ICP method on the Х-7 ICP-MS spectrometer (Thermo Scientific, USA) [29]. The detailed sample decomposition and analytical procedures are described elsewhere [16]. The Hg content was determined in the dry samples using a pyrolyse method on the RA-915+ spectrometer with background correction and a two-chamber atomiser PYRO-915+ (Lumex, Russia) [9, 14].

**Figure 1.** Location of study sites.

116 Trace Elements - Human Health and Environment

#### **3.1. Abundance and distribution of major and trace elements in surface SPM**

The hydrology of the riverine sub-zone of the studied part of the Cai River—Nha Trang Bay estuarine system is strongly influenced by the fill dam. The surface fresh water layer flows seaward over the dam, while the upstream penetration of the near-bottom saline water lenses is blocked [5]. In July 2013, the salinity of the surface fresh-water layer varied from 0% to 36‰. The frontal zone of the contact of fresh and saline waters occurred downstream from the fill dam (sts. 2–3, **Figure 1**), where the horizontal salinity gradient was 3.5‰ per 1 km distance. The temperature (T) of the water column varied in narrow ranges and decreased from the river to the sea from 30 to 29°C in the surface water layer. The pH of the water column increased from neutral in the riverine waters (pH 7) to low-alkaline in the transitional and sea waters (pH 8–9). The dissolved organic carbon (DOC) concentration in the surface water layer varied within the ranges 1.1–2.5 mg l−1 and was distributed uniformly (2.3-2.5 mg l−1) in the frontal zone and transitional waters and exhibited a minimum of 1.1 mg l−1 in the marine part of the transect at salinity >30‰ (**Figure 2**).

The suspended particulate matter (SPM) showed a concentration maximum of 50 mg l−1 near the river mouth (sts. 1) at salinity 0‰ and then a decrease seaward to the values of around 1 mg l−1 at salinities 32–36‰, following from the sedimentation of the coarsest fluvial material in the frontal zone of the estuary at the sharp decrease in the river flow velocity enhanced by the dam influence (**Figure 2**). The distribution pattern of particulate organic carbon (POC) was close to the SPM distribution. The maximal POC concentration (1–1.25 mg l−1) was found in the fluvial part of the estuary (sts. 1–3) at salinities 0–8‰ because of an intensive sedimentation of the organically enriched suspended river material in the frontal zone of the estuary. In the transitional waters, the POC concentration lowers seaward to 0.94 mg l−1at salinities around 20‰ and further to 0.18–0.21 mg l−1 at salinities 32–36‰ (**Figure 2**). The organic carbon content in the SPM (POC, % of dry SPM weight) varied within the range 2–17%. The higher organic carbon content >10% was found in SPM of the marine waters at the salinity >30‰ (**Table 1**).

The distribution of particulate form of Al, Fe, Ti, Li, Zn, Pb, U, Sc, Sn, Bi, Zr, Ba, As, Sr., W, V and Ag followed the distribution of total suspended matter and was characterised by a maximum in the river water and then a sharp decrease seaward of element relative concentration (in μg l−1) with highest horizontal gradients within the salinity interval of 8–20% (**Figure 3**). The absolute concentration of these elements (in μg g−1of the dry SPM weight) followed the same trend of decreasing seaward but was elevated in the both riverine and transitional waters (0–20‰) (**Table 1**, **Figure 3**). The most significant losses of suspended elements occurred in the frontal to transitional zone of the estuary (0–20‰) by an intensive sedimentation of dissolved and suspended river material. The most part of these elements must be supplied to the estuary with the Cai River discharge. In estuaries, the flocculation and coagulation of riverine microcolloids are initiated when the salinity increases. These processes are accompanied by a rapid scavenging of dissolved trace elements from the water column. Further deposition of newly-formed aggregates contributes to the enrichment of sediments in trace elements [32].

The distribution of particulate form of Co, Cu, Ni, Mo and Cr and, in some lesser extent, Mn, Ba, Sn, Sb and Hg is characterised by the most significant loss in the frontal zone of the estuary where the coarsest river material enriched in detrital minerals is deposited at the sharp decrease of the river flow velocity enhanced by the dam (**Table 1**, **Figure 3**). Both relative (in μg l−1) and absolute (in μg g−1 of the dry SPM weight) concentrations of these elements sharply decrease with the highest horizontal gradients at the initial salinity rise (0–8‰). The depletion in these elements in the transitional waters (at salinities 8–32‰) was followed by negligible increase of their relative concentrations and significant increase of

absolute concentrations of Co, Cu, Ni, Mo, and Cr at salinities 32–36‰ (**Table 1**, **Figure 3**). In the stratified Cai River estuary, the significant part of the particulate trace elements may be carried out seaward with the surface water layer. In the marine part of the estuary, with a homogenisation of the water column, most of the fine-grained material of surface water layer enriched in clay minerals, carbonates and trace metals is deposited [5]. Since SPM in

Trace Elements in Suspended Particulate Matter and Sediments of the Cai River: Nha Trang Bay…

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**Figure 2.** SPM, DOC and POC concentrations in surface water layer (in mg l−1).

Trace Elements in Suspended Particulate Matter and Sediments of the Cai River: Nha Trang Bay… http://dx.doi.org/10.5772/intechopen.76471 119

**Figure 2.** SPM, DOC and POC concentrations in surface water layer (in mg l−1).

the river to the sea from 30 to 29°C in the surface water layer. The pH of the water column increased from neutral in the riverine waters (pH 7) to low-alkaline in the transitional and sea waters (pH 8–9). The dissolved organic carbon (DOC) concentration in the surface water layer varied within the ranges 1.1–2.5 mg l−1 and was distributed uniformly (2.3-2.5 mg l−1) in the frontal zone and transitional waters and exhibited a minimum of 1.1 mg l−1 in the marine part

The suspended particulate matter (SPM) showed a concentration maximum of 50 mg l−1 near the river mouth (sts. 1) at salinity 0‰ and then a decrease seaward to the values of around 1 mg l−1 at salinities 32–36‰, following from the sedimentation of the coarsest fluvial material in the frontal zone of the estuary at the sharp decrease in the river flow velocity enhanced by the dam influence (**Figure 2**). The distribution pattern of particulate organic carbon (POC) was close to the SPM distribution. The maximal POC concentration (1–1.25 mg l−1) was found in the fluvial part of the estuary (sts. 1–3) at salinities 0–8‰ because of an intensive sedimentation of the organically enriched suspended river material in the frontal zone of the estuary. In the transitional waters, the POC concentration lowers seaward to 0.94 mg l−1at salinities around 20‰ and further to 0.18–0.21 mg l−1 at salinities 32–36‰ (**Figure 2**). The organic carbon content in the SPM (POC, % of dry SPM weight) varied within the range 2–17%. The higher organic carbon content >10% was found in SPM of the marine waters at the salinity

The distribution of particulate form of Al, Fe, Ti, Li, Zn, Pb, U, Sc, Sn, Bi, Zr, Ba, As, Sr., W, V and Ag followed the distribution of total suspended matter and was characterised by a maximum in the river water and then a sharp decrease seaward of element relative concentration (in μg l−1) with highest horizontal gradients within the salinity interval of 8–20% (**Figure 3**). The absolute concentration of these elements (in μg g−1of the dry SPM weight) followed the same trend of decreasing seaward but was elevated in the both riverine and transitional waters (0–20‰) (**Table 1**, **Figure 3**). The most significant losses of suspended elements occurred in the frontal to transitional zone of the estuary (0–20‰) by an intensive sedimentation of dissolved and suspended river material. The most part of these elements must be supplied to the estuary with the Cai River discharge. In estuaries, the flocculation and coagulation of riverine microcolloids are initiated when the salinity increases. These processes are accompanied by a rapid scavenging of dissolved trace elements from the water column. Further deposition of newly-formed aggregates contributes to the enrichment of

The distribution of particulate form of Co, Cu, Ni, Mo and Cr and, in some lesser extent, Mn, Ba, Sn, Sb and Hg is characterised by the most significant loss in the frontal zone of the estuary where the coarsest river material enriched in detrital minerals is deposited at the sharp decrease of the river flow velocity enhanced by the dam (**Table 1**, **Figure 3**). Both relative (in μg l−1) and absolute (in μg g−1 of the dry SPM weight) concentrations of these elements sharply decrease with the highest horizontal gradients at the initial salinity rise (0–8‰). The depletion in these elements in the transitional waters (at salinities 8–32‰) was followed by negligible increase of their relative concentrations and significant increase of

of the transect at salinity >30‰ (**Figure 2**).

118 Trace Elements - Human Health and Environment

>30‰ (**Table 1**).

sediments in trace elements [32].

absolute concentrations of Co, Cu, Ni, Mo, and Cr at salinities 32–36‰ (**Table 1**, **Figure 3**). In the stratified Cai River estuary, the significant part of the particulate trace elements may be carried out seaward with the surface water layer. In the marine part of the estuary, with a homogenisation of the water column, most of the fine-grained material of surface water layer enriched in clay minerals, carbonates and trace metals is deposited [5]. Since SPM in


**3.2. Abundance and distribution of major and trace elements in surface sediments**

sediment, with the highest clay content, is from marine stations 7–8.

**Figure 3.** Major and trace elements contents in SPM (in μg l−1).

extent) increase seaward in the sediments with the clay-sized materials.

The percent of the content of sand- (63 μm–2 mm), silt- (2–63 μm) and clay- (<2 μm) sized material in the studied sediments is shown in **Table 2**. The sediments near the river mouth were mostly sandy, while the coarsest sediment is from station 4. Downstream, in the transitional sub-zone, the sediment contains less sand and more silt and clay. The most fine-grained

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The total contents of Al, Fe, Ca, Mn, Ti, TOC and TIC are reported in **Table 3**. The mean contents of the major elements are within the range of the Clark contents in shale, pelagic clays and average world riverbed sediments [33–37]. The distribution of Fe and Ti in the river-sea transect is similar to that of Al. The observed distribution of major elements in the sediments illustrates the grain size and mineral fractionation processes. Al, Fe, Ti and Mn (to a lesser

The distribution of the inorganic carbon (TIC) content in the sediments along the river-sea transect is characterised by only two significant values in the marine part (2,1 % at st. 7 and

**Table 1.** Major and trace elements contents in SPM (in μg g−1, except for Al, Fe, Ti, Mn, Ca and POC in % of dry weight).

transitional and marine waters were enriched in organic carbon, the particulate organic matter of terrigenous and/or planktonogenous origin most probably contributed to trace element accumulation in SPM in the bay.

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**Figure 3.** Major and trace elements contents in SPM (in μg l−1).

transitional and marine waters were enriched in organic carbon, the particulate organic matter of terrigenous and/or planktonogenous origin most probably contributed to trace element

**Table 1.** Major and trace elements contents in SPM (in μg g−1, except for Al, Fe, Ti, Mn, Ca and POC in % of dry weight).

**Station (S ‰) 1 (0‰) 3 (8‰) 4 (18‰) 7 (32‰) 8 (36‰)** Li 57.09 60.32 57.37 9.50 8.67 Al 15.11 15.41 14.61 1.30 0.78 Ca 0.31 0.16 0.16 0.83 0.82 Sc 15.85 15.42 16.17 5.65 3.52 Ti 0.18 0.15 0.12 0.03 0.02 V 94.07 109.59 127.60 63.59 17.39 Cr 187.21 57.34 72.09 66.18 91.86 Mn 737.05 410.08 342.46 461.65 479.05 Fe 4.67 4.77 5.31 0.45 0.25 Co 175.41 13.66 12.24 60.91 116.96 Ni 384.24 52.48 61.56 119.56 171.20 Cu 201.66 31.60 34.25 119.16 201.11 Zn 116.07 119.56 103.54 28.00 35.7 As 42.42 34.88 39.85 6.61 89.64 Sr 45.32 46.63 61.39 201.82 201.40 Zr 27.46 22.74 22.62 4.25 2.52 Mo 51.58 9.13 9.45 10.97 13.57 Ag 0.22 0.18 0.22 0.25 0.13 Sn 9.19 8.56 6.57 2.53 1.77 Sb 0.97 0.93 0.94 0.22 0.20 Cs 16.30 15.12 11.80 0.97 0.77 Ba 286.38 205.25 248.46 28.48 157.93 W 12.01 5.63 5.43 1.64 1.06 Hg 0.89 0.49 1.11 2.28 0.91 Pb 65.17 63.05 75.28 44.76 26.88 Bi 9.66 9.84 14.59 1.14 0.42 U 9.08 10.33 8.77 0.49 0.43 POC 3.68 3.85 9.39 7.87 7.29

120 Trace Elements - Human Health and Environment

accumulation in SPM in the bay.

#### **3.2. Abundance and distribution of major and trace elements in surface sediments**

The percent of the content of sand- (63 μm–2 mm), silt- (2–63 μm) and clay- (<2 μm) sized material in the studied sediments is shown in **Table 2**. The sediments near the river mouth were mostly sandy, while the coarsest sediment is from station 4. Downstream, in the transitional sub-zone, the sediment contains less sand and more silt and clay. The most fine-grained sediment, with the highest clay content, is from marine stations 7–8.

The total contents of Al, Fe, Ca, Mn, Ti, TOC and TIC are reported in **Table 3**. The mean contents of the major elements are within the range of the Clark contents in shale, pelagic clays and average world riverbed sediments [33–37]. The distribution of Fe and Ti in the river-sea transect is similar to that of Al. The observed distribution of major elements in the sediments illustrates the grain size and mineral fractionation processes. Al, Fe, Ti and Mn (to a lesser extent) increase seaward in the sediments with the clay-sized materials.

The distribution of the inorganic carbon (TIC) content in the sediments along the river-sea transect is characterised by only two significant values in the marine part (2,1 % at st. 7 and


**Table 2.** Sand, silt, clay, TOC and TIC contents in sediments (in % of dry weight).

0.88% at st. 8). Sedimentary organic carbon (TOC) varied within the range of 0.6–1.8% and showed no affinity to the other major or trace elements that were studied. This may be due to the intensive microbial decomposition of particulate organic matter, which occurs in the water column during estuarine sedimentation processes [38, 39]. The post-depositional diagenetic reactions, which are enhanced by resuspension processes at sediment disturbance events (such as tides, storms and upwelling), may also contribute to a destruction of sedimentary organic matter and the formation of organic-poor sediments [40].

The mean content of the major part of the studied trace elements (Li, V, Cr, Co, Ni, Cu, Zn, As, Sr, Zr, Mo, Cd, Sn, Sb, Cs, Ba, Hg and Pb) in the sediments from the Cai River estuary and Nha Trang Bay is lower or corresponds to the reference values for shale, pelagic clays and the average world riverbed sediments (**Table 3**). The Ag content was negligible or below the detection limit at all locations along the salinity gradient. Thus, natural enrichment of Ag reported in the previous study [19] had a temporary/impact character [33–37]. However, relative sediment enrichment with Bi, W and, at some sites, with Sr. needs special study.

To normalise the obtained geochemical data for the grain-size effects and identify the enrichment zones along the salinity gradient, the metal/Al ratios were calculated [27, 41, 42]. The distribution of metal/Al ratios along the Cai River—Nha Trang Bay transect is provided in **Figure 4**. The results revealed associations of elements that are characterised by a similar geochemical behaviour in the sediments along the salinity gradient. Sedimentary Fe, Ti, Li, Sc, Co, Cs, Zr, Cr, Zn, Co, Ni, Cu, Pb, Sn, V As, U and Mo varied in relatively narrow ranges. Major part of these elements tended to increase seaward with an elevation at station 7 at heightened carbonate content. The observed distribution of the normalised trace element contents reflects the association with and/or inclusion of Fe, Ti, Li and trace elements in the lattices of clay minerals that constitute the bulk of the fine-grained sedimentary material accumulated in the sea floor depression in the bay [42]. Sedimentary Bi and W decreased significantly from river to the sea. These elements may be associated with the coarsest river material enriched in detrital minerals which is mostly deposited in the riverine part of the estuary. The distribution of Sr. and Ca and, in a lesser extent, of Mn and Ba is largely controlled by the total inorganic carbon (TIC) content in the sediments. These elements form low-soluble carbonates in aquatic environments [36]. The distribution of trace elements in

sediments is strongly influenced by the water column stratification because of the natural fractionation and deposition of materials of different grain sizes at sites, which are deter-

**Table 3.** Major and trace elements contents in sediments (in μg g−1, except for Al, Fe, Ti, Mn, Ca, TOC and TIC in % of

**Element Mean SD Range Shalea Pelagic clay<sup>a</sup> River sedb** Al 10.8 2.05 7.67–12.23 8.8 8.4 4.3 Fe 3.98 0.64 2.46–4.51 4.72 6.5 2.5 Ti 0.36 0.05 0.25–0.40 0.46 0.46 0.31 Mn 0.04 0.01 0.03–0.06 0.085 0.67 0.05 Ca 1.26 1.93 0.36–6.16 1.6 1.0 1.7 Li 47.9 8.2 34.7–62.7 66 57 20 V 89.3 12.0 61.4–100 130 120 50 Cr 45.8 9.9 27.9–66.7 90 90 50 Co 8.5 1.8 4.9–12.2 19 74 15 Ni 23.2 6.12 14.2–38.1 50 230 25 Cu 18.5 4.12 12.1–26.9 45 250 20 Zn 104.6 16.17 69.8–121 95 170 60 As 22.2 6.3 12.2–30.9 13 20 6 Sr 145 166 61.3–605 170 180 150 Zr 85.2 10.4 61.7–104 160 150 250 Mo 3.0 1.5 0.6–5.1 2.6 27 1.5 Ag 0.1 0.01 0.07–0.1 0.07 0.11 0.1 Sn 5.9 1.3 3.7–7.9 3.0 4.0 4.0 Sb 1.2 0.2 0.9–1.4 1.5 1.0 2.0 Cs 11.1 1.6 7.9–13.1 5.0 6.0 4.0 Ba 274 18 256–323 580 2300 290 Hg 0.03 0.01 0.02–0.05 0.18 0.10 0.05 Pb 54.4 11.0 35.1–61.6 20 80 15 Bi 9.5 9.0 0.9–27.6 0.43 0.53 0.2 U 6.6 1.5 3.9–8.0 2.7 2.6 3 W 10.4 7.2 2.4–24.6 0.3 0.42 0.4 TOC 1.36 0.39 0.61–1.77 — — 1.4 TIC 1.49 0.69 <0.01–2.10 — — 0.4

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mined by hydrodynamic conditions [5, 13, 43, 44].

a

b

Cited from [24].

Cited from [46].

dry weight).

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Cited from [24].

0.88% at st. 8). Sedimentary organic carbon (TOC) varied within the range of 0.6–1.8% and showed no affinity to the other major or trace elements that were studied. This may be due to the intensive microbial decomposition of particulate organic matter, which occurs in the water column during estuarine sedimentation processes [38, 39]. The post-depositional diagenetic reactions, which are enhanced by resuspension processes at sediment disturbance events (such as tides, storms and upwelling), may also contribute to a destruction of sedimen-

**Stations Sand (63 μm–2 mm) Silt (2–63 μm) Clay (<2 μm) TOC TIC** 49.58 23.96 26.46 1.77 0 47.81 28.28 23.91 1.49 0 71.03 19.23 9.74 0.89 0 28.47 47.48 24.05 1.55 0 49.31 37.6 13.09 1.76 0 3.26 42.69 54.05 0.61 2.11 2.37 28.96 68.67 1.02 0.88

The mean content of the major part of the studied trace elements (Li, V, Cr, Co, Ni, Cu, Zn, As, Sr, Zr, Mo, Cd, Sn, Sb, Cs, Ba, Hg and Pb) in the sediments from the Cai River estuary and Nha Trang Bay is lower or corresponds to the reference values for shale, pelagic clays and the average world riverbed sediments (**Table 3**). The Ag content was negligible or below the detection limit at all locations along the salinity gradient. Thus, natural enrichment of Ag reported in the previous study [19] had a temporary/impact character [33–37]. However, relative sediment enrichment with Bi, W and, at some sites, with Sr. needs spe-

To normalise the obtained geochemical data for the grain-size effects and identify the enrichment zones along the salinity gradient, the metal/Al ratios were calculated [27, 41, 42]. The distribution of metal/Al ratios along the Cai River—Nha Trang Bay transect is provided in **Figure 4**. The results revealed associations of elements that are characterised by a similar geochemical behaviour in the sediments along the salinity gradient. Sedimentary Fe, Ti, Li, Sc, Co, Cs, Zr, Cr, Zn, Co, Ni, Cu, Pb, Sn, V As, U and Mo varied in relatively narrow ranges. Major part of these elements tended to increase seaward with an elevation at station 7 at heightened carbonate content. The observed distribution of the normalised trace element contents reflects the association with and/or inclusion of Fe, Ti, Li and trace elements in the lattices of clay minerals that constitute the bulk of the fine-grained sedimentary material accumulated in the sea floor depression in the bay [42]. Sedimentary Bi and W decreased significantly from river to the sea. These elements may be associated with the coarsest river material enriched in detrital minerals which is mostly deposited in the riverine part of the estuary. The distribution of Sr. and Ca and, in a lesser extent, of Mn and Ba is largely controlled by the total inorganic carbon (TIC) content in the sediments. These elements form low-soluble carbonates in aquatic environments [36]. The distribution of trace elements in

tary organic matter and the formation of organic-poor sediments [40].

**Table 2.** Sand, silt, clay, TOC and TIC contents in sediments (in % of dry weight).

122 Trace Elements - Human Health and Environment

cial study.

b Cited from [46].

**Table 3.** Major and trace elements contents in sediments (in μg g−1, except for Al, Fe, Ti, Mn, Ca, TOC and TIC in % of dry weight).

sediments is strongly influenced by the water column stratification because of the natural fractionation and deposition of materials of different grain sizes at sites, which are determined by hydrodynamic conditions [5, 13, 43, 44].

highest in sediments in the frontal zone (sts. 2–3). The percent content of the weak-acid-soluble Fe, which is mostly comprised of easily soluble amorphous oxides, was constant in the transitional zone (10–11% of the total content), reached the maximum of 18% in the coarsest sediment from station 4 and lowered to 7–8% in the bay sediments. The percent content of the pyrophosphate-soluble Fe was low and decreased from 6–1% of the total Fe along the salinity gradient. The total and percent contents of oxalate-soluble Mn (18–58%), weak-acid-soluble Mn (13–54%) and pyrophosphate-soluble Mn (9–28%) were the highest in the coarse sediments of the frontal zone (sts. 2–3). Seaward, the contents of the studied forms of Mn decreased in the sediments in the transitional sub-zone (sts. 4–6) and increased again in the bay (sts. 7–8). The oxalate-soluble form comprised 9–14% (12% on average) of the total content for Cr, 3–11% (8%) for Pb, 5–11% (7%) for Zn, 3–23%(15%) for Cu, 5–12% (9%) for Ni and 13–22% (17%) for Co. The weak-acid-soluble form comprised 8–12% (9% on average) of the total content for Cr, 23–34% (30%) for Pb, 4–19% (12%) for Zn, 1–8% (2%) for Cu, 8–24% (17%) for Ni and 20–66% (44%) for Co. The contents of pyrophosphate-soluble form were below the detection limit for Ni and Co. This form comprised 0.3–3% (2% on average) of the total content for Cr, 5–17% (11%) for Pb, 1–16% (10%) for Zn and 1–26% (12%) for Cu. According to the comparative extractability from sediments, Ni, Zn, Cr and Cu are low-labile and mainly occur in the residual phase. These metals were mainly extracted in the detrital fraction, which emphasises the importance of natural weathering and erosion in drainage basins. Fe and Zn are moderately labile and occur in the less resistant phases such as crystallised Fe/Mn oxides and organic compounds that may be a threat in the long term. Mn, Co and Pb are labile, held in ion exchange positions, bound to easily soluble amorphous Fe/Mn compounds and weakly held in organic matter. The high levels of acid-soluble Pb and Co (30 and 43% of the total content on average, respectively) compared to previously studied estuarine and coastal sediments contributes to a contamination problem in the Nha Trang Bay, which arises from the Cai River discharges, while the elevated level of easily reducible and organically bound Pb fractions (8 and 11% of the total content on average, respectively) also contributes to the anthropogenic input of Pb [5, 13].

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The contents of oxalate-soluble (amorphous) forms were higher than the contents of pyrophosphate-soluble (organically bound) forms at all sites for Fe, Mn, Ni, Co and Cr (**Table 4**). Therefore, the most bioavailable parts of Ni, Co and Cr are bound to amorphous Fe and Mn oxyhydroxides and acid-soluble organic compounds. The contents of pyrophosphate-soluble forms were higher than the contents of oxalate-soluble forms at most of the sites for Pb, Zn and at some sites for Cu. Among the elements studied, most of the bioavailable Pb, Zn and Cu was most likely bound to organic substances. According to the mean determined amounts of the oxalate-soluble, pyrophosphate-soluble and weak-acid-soluble forms, the studied elements can be arranged in the following increasing order of average potential bioavailability: Cr < Ni < Cu < Zn < Fe < <Pb < Co < <Mn. The most bioavailable trace elements in sediments that were studied were scavenged by amorphous iron oxyhydroxides in the course of estuarine sedimentation. This result supports the fact that Fe and Mn oxyhydroxides largely con-

**Figure 5** illustrates the distribution of the ecologically most significant weak-acid-soluble (labile) fraction along the river-sea transect. Mn, Co and Pb have the highest percent contents of the labile form but exhibit different spatial distributions showing some sporadic enrichments

trol the bioavailability in sediments [7, 19, 32].

**Figure 4.** Element/Al ratio in sediments.

#### **3.3. Speciation of major and trace elements in surface sediments**

Major and trace elements are bound to a variety of sediment fractions that range from easily extractable (and bioavailable) to resistant residual mineral phases [45–49]. The total contents of the oxalate-soluble, pyrophosphate-soluble and weak-acid-soluble forms of Fe, Mn, Cr, Zn, Cu, Pb, Ni and Co are provided in **Table 4**. In this work, ammonium oxalate (pH 3.2–3.3) served to mobilise the easily soluble amorphous Fe-oxyhydroxides and acid-soluble fulvates [30, 31, 50]. Sodium pyrophosphate (pH 10) was used to remove organically bound metals from the sediments. This extract also mobilised part of easily soluble amorphous Fe-oxyhydroxides [13]. Acetic acid removed the labile metals in ion exchange positions, the easily soluble amorphous compounds of iron and manganese, the carbonates and the metals that are weakly held in organic matter [27].

The total content of the oxalate-soluble (amorphous) Fe increased from the river to the sea, whereas its percent content varied insignificantly (18–24% of the total content) and is the highest in sediments in the frontal zone (sts. 2–3). The percent content of the weak-acid-soluble Fe, which is mostly comprised of easily soluble amorphous oxides, was constant in the transitional zone (10–11% of the total content), reached the maximum of 18% in the coarsest sediment from station 4 and lowered to 7–8% in the bay sediments. The percent content of the pyrophosphate-soluble Fe was low and decreased from 6–1% of the total Fe along the salinity gradient. The total and percent contents of oxalate-soluble Mn (18–58%), weak-acid-soluble Mn (13–54%) and pyrophosphate-soluble Mn (9–28%) were the highest in the coarse sediments of the frontal zone (sts. 2–3). Seaward, the contents of the studied forms of Mn decreased in the sediments in the transitional sub-zone (sts. 4–6) and increased again in the bay (sts. 7–8).

The oxalate-soluble form comprised 9–14% (12% on average) of the total content for Cr, 3–11% (8%) for Pb, 5–11% (7%) for Zn, 3–23%(15%) for Cu, 5–12% (9%) for Ni and 13–22% (17%) for Co. The weak-acid-soluble form comprised 8–12% (9% on average) of the total content for Cr, 23–34% (30%) for Pb, 4–19% (12%) for Zn, 1–8% (2%) for Cu, 8–24% (17%) for Ni and 20–66% (44%) for Co. The contents of pyrophosphate-soluble form were below the detection limit for Ni and Co. This form comprised 0.3–3% (2% on average) of the total content for Cr, 5–17% (11%) for Pb, 1–16% (10%) for Zn and 1–26% (12%) for Cu. According to the comparative extractability from sediments, Ni, Zn, Cr and Cu are low-labile and mainly occur in the residual phase. These metals were mainly extracted in the detrital fraction, which emphasises the importance of natural weathering and erosion in drainage basins. Fe and Zn are moderately labile and occur in the less resistant phases such as crystallised Fe/Mn oxides and organic compounds that may be a threat in the long term. Mn, Co and Pb are labile, held in ion exchange positions, bound to easily soluble amorphous Fe/Mn compounds and weakly held in organic matter. The high levels of acid-soluble Pb and Co (30 and 43% of the total content on average, respectively) compared to previously studied estuarine and coastal sediments contributes to a contamination problem in the Nha Trang Bay, which arises from the Cai River discharges, while the elevated level of easily reducible and organically bound Pb fractions (8 and 11% of the total content on average, respectively) also contributes to the anthropogenic input of Pb [5, 13].

The contents of oxalate-soluble (amorphous) forms were higher than the contents of pyrophosphate-soluble (organically bound) forms at all sites for Fe, Mn, Ni, Co and Cr (**Table 4**). Therefore, the most bioavailable parts of Ni, Co and Cr are bound to amorphous Fe and Mn oxyhydroxides and acid-soluble organic compounds. The contents of pyrophosphate-soluble forms were higher than the contents of oxalate-soluble forms at most of the sites for Pb, Zn and at some sites for Cu. Among the elements studied, most of the bioavailable Pb, Zn and Cu was most likely bound to organic substances. According to the mean determined amounts of the oxalate-soluble, pyrophosphate-soluble and weak-acid-soluble forms, the studied elements can be arranged in the following increasing order of average potential bioavailability: Cr < Ni < Cu < Zn < Fe < <Pb < Co < <Mn. The most bioavailable trace elements in sediments that were studied were scavenged by amorphous iron oxyhydroxides in the course of estuarine sedimentation. This result supports the fact that Fe and Mn oxyhydroxides largely control the bioavailability in sediments [7, 19, 32].

**3.3. Speciation of major and trace elements in surface sediments**

that are weakly held in organic matter [27].

**Figure 4.** Element/Al ratio in sediments.

124 Trace Elements - Human Health and Environment

Major and trace elements are bound to a variety of sediment fractions that range from easily extractable (and bioavailable) to resistant residual mineral phases [45–49]. The total contents of the oxalate-soluble, pyrophosphate-soluble and weak-acid-soluble forms of Fe, Mn, Cr, Zn, Cu, Pb, Ni and Co are provided in **Table 4**. In this work, ammonium oxalate (pH 3.2–3.3) served to mobilise the easily soluble amorphous Fe-oxyhydroxides and acid-soluble fulvates [30, 31, 50]. Sodium pyrophosphate (pH 10) was used to remove organically bound metals from the sediments. This extract also mobilised part of easily soluble amorphous Fe-oxyhydroxides [13]. Acetic acid removed the labile metals in ion exchange positions, the easily soluble amorphous compounds of iron and manganese, the carbonates and the metals

The total content of the oxalate-soluble (amorphous) Fe increased from the river to the sea, whereas its percent content varied insignificantly (18–24% of the total content) and is the **Figure 5** illustrates the distribution of the ecologically most significant weak-acid-soluble (labile) fraction along the river-sea transect. Mn, Co and Pb have the highest percent contents of the labile form but exhibit different spatial distributions showing some sporadic enrichments


along the salinity gradient. Thus, the distribution of the most abundant labile Co is complicated by a pronounced maximum of 67% in the sediment at station 5. The sediments are mostly enriched with labile Fe, Zn, Cr and Ni in the frontal and transitional sub-zones (sts. 2–5). Cu exhibit the lowest contents of the labile form. Therefore, in the studied sediments, Cu is most

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likely bound to the residual mineral phase that is comprised of detrital heavy minerals.

The suspended particulate matter (SPM) showed a concentration maximum (50 mg l

salinities 32–36‰, following from the sedimentation of the coarsest fluvial material in the frontal zone of the estuary at the sharp decrease in the river flow velocity enhanced by the dam influence. The distribution pattern of particulate organic carbon (POC) was close to

tent in the SPM (POC, % of dry SPM weight) varied within the range 2–17%. The higher organic carbon content >10% was found in SPM of the marine waters at the salinity >30‰. The distribution of particulate form of Al, Fe, Ti, Li, Zn, Pb, U, Sc, Sn, Bi, Zr, Ba, As, Sr., W, V and Ag followed the distribution of total suspended matter and was characterised by a maximum in the river water and then a sharp decrease seaward of element relative concentration (in μg l−1) with highest horizontal gradients within the salinity interval of 8–20%. The most part of these elements must be supplied to the estuary with the Cai River discharge. The distribution of particulate form of Co, Cu, Ni, Mo and Cr and, in a lesser extent, Mn, Ba, Sn, Sb and Hg is characterised by the most significant loss in the frontal zone of the estuary where the coarsest river material enriched in detrital minerals and pronounced increase of their absolute concentrations

near the river mouth and then a decrease seaward to the values of around 1 mg l

the SPM distribution and varied within the range 0.18–1.25 mg l

**Figure 5.** Weak acid-soluble (labile) metal form in sediments.

−1)

−1 at

−1. The organic carbon con-

**4. Conclusions**

**Table 4.** Major and trace element form contents in sediments (in μg g−1).

**Figure 5.** Weak acid-soluble (labile) metal form in sediments.

along the salinity gradient. Thus, the distribution of the most abundant labile Co is complicated by a pronounced maximum of 67% in the sediment at station 5. The sediments are mostly enriched with labile Fe, Zn, Cr and Ni in the frontal and transitional sub-zones (sts. 2–5). Cu exhibit the lowest contents of the labile form. Therefore, in the studied sediments, Cu is most likely bound to the residual mineral phase that is comprised of detrital heavy minerals.

## **4. Conclusions**

**Stations Cu Zn Ni Co Pb Cr Fe Mn**

 0.4 18 5.6 4 16.4 3.4 4540 344 0.4 14 4.8 4.6 14 4.8 5620 144 0.2 13.2 3.4 2.2 13.4 3.2 2840 36 0.2 14 3 6 16 4.8 7640 58 ≤0.2 4 2.4 1.8 12 5.2 2660 164 0.2 15.8 3 3.4 20 3.4 3440 52 ≤0.2 7.6 5 2.6 15.6 5.6 2920 264 Mean 0.525 12.9 3.82 3.6 15.84 4.22 4244 133.6

 5.6 18.6 ≤0.2 ≤0.2 4.8 1 2580 180 2 10.4 ≤0.2 ≤0.2 7 0.8 2940 82 2.6 9 ≤0.2 ≤0.2 7 ≤0.2 1300 30 0.8 14.8 ≤0.2 ≤0.2 5.4 0.4 1480 32 0.4 1.2 ≤0.2 ≤0.2 5 0.8 420 50 1 13.4 ≤0.2 ≤0.2 9.6 1.6 1360 42 0.2 2 ≤0.2 ≤0.2 7.6 ≤0.2 600 90 Mean 2.24 10.84 ≤0.2 ≤0.2 5.74 1 1652 65.9

 3 5.6 1.4 1.6 5.4 5.3 11,300 375 3.7 7 2.2 1 6.8 5.2 10,550 172 2.5 8 1.6 0.6 3.4 4 5480 75 2.1 6.8 1.3 2.2 5.4 6 8250 87 0.4 6.5 2.2 1.2 1.1 7 7500 158 4.4 8.7 2.6 1.5 4.1 5 7550 90 0.5 8.1 2 1.9 1.2 6.2 8430 255 Mean 2.93 7.42 2 1.44 4.56 5.57 8572 152.1

2.8 12.3 16.5 42.4 29.1 9.2 10.6 32.0

12.1 10.4 — — 10.6 2.2 4.1 15.8

15.8 7.1 8.6 16.9 8.4 12.2 21.5 36.4

Weak acid-soluble

126 Trace Elements - Human Health and Environment

Mean

Mean

Mean

(% of total content)

**Table 4.** Major and trace element form contents in sediments (in μg g−1).

(% of total content)

Oxalate-soluble

(% of total content)

Pyrophosphate-soluble

The suspended particulate matter (SPM) showed a concentration maximum (50 mg l −1) near the river mouth and then a decrease seaward to the values of around 1 mg l −1 at salinities 32–36‰, following from the sedimentation of the coarsest fluvial material in the frontal zone of the estuary at the sharp decrease in the river flow velocity enhanced by the dam influence. The distribution pattern of particulate organic carbon (POC) was close to the SPM distribution and varied within the range 0.18–1.25 mg l −1. The organic carbon content in the SPM (POC, % of dry SPM weight) varied within the range 2–17%. The higher organic carbon content >10% was found in SPM of the marine waters at the salinity >30‰.

The distribution of particulate form of Al, Fe, Ti, Li, Zn, Pb, U, Sc, Sn, Bi, Zr, Ba, As, Sr., W, V and Ag followed the distribution of total suspended matter and was characterised by a maximum in the river water and then a sharp decrease seaward of element relative concentration (in μg l−1) with highest horizontal gradients within the salinity interval of 8–20%. The most part of these elements must be supplied to the estuary with the Cai River discharge. The distribution of particulate form of Co, Cu, Ni, Mo and Cr and, in a lesser extent, Mn, Ba, Sn, Sb and Hg is characterised by the most significant loss in the frontal zone of the estuary where the coarsest river material enriched in detrital minerals and pronounced increase of their absolute concentrations at salinities 32–36‰. In the stratified Cai River estuary, the significant part of the particulate trace elements may be carried out seaward with the surface water layer. In the marine part of the estuary, with a homogenisation of the water column, most of the fine-grained material of surface water layer enriched in organic matter and trace metals is deposited.

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Sedimentary Fe, Ti, Li, Sc, Co, Cs, Zr, Cr, Zn, Co, Ni, Cu, Pb, Sn, V As, U and Mo varied in relatively narrow ranges along the salinity gradient and tend to increase seaward. These elements are most likely controlled by the accumulation of their most fine-grained aluminosilicate host minerals and materials in the sea floor depression of the marine sub-zone. Sedimentary Bi and W, are generally uniformly low but tend to decrease seaward. These elements may be associated with the coarsest river material enriched in detrital minerals which is mostly deposited in the riverine part of the estuary. The distribution of Sr. and Ca and, in a lesser extent, of Mn and Ba is largely controlled by the total inorganic carbon (TIC) content in the sediments.

The distribution of trace elements in SPM and sediments of Cai River—Nha Trang Bay estuarine system is strongly influenced by the water column stratification because of the natural fractionation and deposition of materials of different grain sizes at sites, which are determined by hydrodynamic conditions.

Assuming that the mean determined amounts of the oxalate-soluble, pyrophosphate-soluble and weak-acid-soluble forms are a measure of the potential metal bioavailability in sediments of the Cai River—Nha Trang Bay estuarine system, the studied elements can be arranged in the following increasing order of average potential bioavailability: Cr < Ni < Cu < Zn < Fe < <Pb < Co < <Mn. This sequence is true for sediments in different sub-zones of the water-mixing zone: estuary (transitional waters) and sea (bay). Metal form study revealed the highest percent contents of the labile (weak acid-soluble) form for Mn, Co and Pb in the sediments. The high levels of labile Pb and Co (30 and 43% of the total content in sediment, on average, respectively) contribute to a heavy metal contamination problem in the Nha Trang Bay, which arises from the Cai River discharge. The elevated level of amorphous (oxalate-soluble) and organically bound (pyrophosphate-soluble) Pb fractions (8 and 11% of the total content in sediment, on average, respectively) also contribute to the anthropogenic input of Pb. The most bioavailable parts of the studied trace metals are associated with easily soluble amorphous Fe and Mn oxyhydroxides.

## **Acknowledgements**

This research was performed in the framework of the state assignment of FASO Russia (theme No. 0149-2018-0005).

## **Author details**

Sofia Koukina\* and Nikolay Lobus

\*Address all correspondence to: skoukina@gmail.com

Shirshov Institute of Oceanology of RAS, Russia

## **References**

at salinities 32–36‰. In the stratified Cai River estuary, the significant part of the particulate trace elements may be carried out seaward with the surface water layer. In the marine part of the estuary, with a homogenisation of the water column, most of the fine-grained material of

Sedimentary Fe, Ti, Li, Sc, Co, Cs, Zr, Cr, Zn, Co, Ni, Cu, Pb, Sn, V As, U and Mo varied in relatively narrow ranges along the salinity gradient and tend to increase seaward. These elements are most likely controlled by the accumulation of their most fine-grained aluminosilicate host minerals and materials in the sea floor depression of the marine sub-zone. Sedimentary Bi and W, are generally uniformly low but tend to decrease seaward. These elements may be associated with the coarsest river material enriched in detrital minerals which is mostly deposited in the riverine part of the estuary. The distribution of Sr. and Ca and, in a lesser extent, of Mn and Ba is largely controlled by the total inorganic carbon (TIC) content in the sediments.

The distribution of trace elements in SPM and sediments of Cai River—Nha Trang Bay estuarine system is strongly influenced by the water column stratification because of the natural fractionation and deposition of materials of different grain sizes at sites, which are deter-

Assuming that the mean determined amounts of the oxalate-soluble, pyrophosphate-soluble and weak-acid-soluble forms are a measure of the potential metal bioavailability in sediments of the Cai River—Nha Trang Bay estuarine system, the studied elements can be arranged in the following increasing order of average potential bioavailability: Cr < Ni < Cu < Zn < Fe < <Pb < Co < <Mn. This sequence is true for sediments in different sub-zones of the water-mixing zone: estuary (transitional waters) and sea (bay). Metal form study revealed the highest percent contents of the labile (weak acid-soluble) form for Mn, Co and Pb in the sediments. The high levels of labile Pb and Co (30 and 43% of the total content in sediment, on average, respectively) contribute to a heavy metal contamination problem in the Nha Trang Bay, which arises from the Cai River discharge. The elevated level of amorphous (oxalate-soluble) and organically bound (pyrophosphate-soluble) Pb fractions (8 and 11% of the total content in sediment, on average, respectively) also contribute to the anthropogenic input of Pb. The most bioavailable parts of the studied trace metals are associated with easily soluble amorphous Fe and Mn oxyhydroxides.

This research was performed in the framework of the state assignment of FASO Russia (theme

surface water layer enriched in organic matter and trace metals is deposited.

mined by hydrodynamic conditions.

128 Trace Elements - Human Health and Environment

**Acknowledgements**

No. 0149-2018-0005).

**Author details**

Sofia Koukina\* and Nikolay Lobus

\*Address all correspondence to: skoukina@gmail.com

Shirshov Institute of Oceanology of RAS, Russia


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**Chapter 7**

**Provisional chapter**

**A Simple and Highly Structured Procaine**

**Determination of Mercury Species in Water**

Dyab A. Al-Eryani, Waqas Ahmad, Zeinab M. Saigl,

Dyab A. Al-Eryani, Waqas Ahmad, Zeinab M. Saigl, Hassan Alwael, Saleh O. Bahaffi, Yousry M. Moustafa

Yousry M. Moustafa and Mohammad S. El-Shahawi

Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

Hassan Alwael, Saleh O. Bahaffi,

and Mohammad S. El-Shahawi

http://dx.doi.org/10.5772/intechopen.73397

probe procaine hydrochloride (PQ+

of the ternary ion associate complex [(PQ+

quenching, ternary ion associate complex

**Abstract**

**1. Introduction**

**Hydrochloride as Fluorescent Quenching Chemosensor**

An ultrasensitive, simple and highly selective spectrofluorometric strategy for quantifying traces of mercury(II) in environmental water has been established using the fluorescent

> ) 2 .(HgI4 )

**Keywords:** spectrofluorometry, mercury(II), fluorescence, procaine hydrochloride,

iodide media at pH 9.0–10.0 with its subsequent extraction onto dichloromethane accompanied by a change in fluorescence intensity at λex/em = 268/333 nm. The developed strategy exhibited a linear range of 1–114 μg L−1 with lower limit of detection (LOD) and quantification (LOQ) of mercury(II) 1.3 and 3.98 nM, respectively. Intra and inter-day laboratory accuracy and precision for trace analysis of mercury(II) in water were performed. Complexed mercury(II) in environmental water, chemical speciation and successful literature comparison was performed. The proposed system offered excellent selectivity towards mercury(II) ions examined in the presence of competent ions in excess, relevant to real water samples. The method was applied for analysis of mercury(II) in tap water samples. Statistical comparison (Student's *t* and *F* tests) of the proposed method with the reference ICP-OES method revealed no significant differences in the accuracy and precision.

). The procedure was based upon the formation

.Cl<sup>−</sup>

and mercury(II) in

2−] between PQ+

.Cl<sup>−</sup>

**A Simple and Highly Structured Procaine Hydrochloride** 

DOI: 10.5772/intechopen.73397

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution,

© 2018 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use,

distribution, and reproduction in any medium, provided the original work is properly cited.

and reproduction in any medium, provided the original work is properly cited.

Heavy metal pollution is global level environmental concern, which poses serious implications towards human health [1, 2]. Heavy metal pollution has attained considerable interest

**for Trace Determination of Mercury Species in Water**

**as Fluorescent Quenching Chemosensor for Trace** 


**Provisional chapter**

## **A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor for Trace Determination of Mercury Species in Water A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor for Trace Determination of Mercury Species in Water**

DOI: 10.5772/intechopen.73397

Dyab A. Al-Eryani, Waqas Ahmad, Zeinab M. Saigl, Hassan Alwael, Saleh O. Bahaffi, Yousry M. Moustafa and Mohammad S. El-Shahawi Hassan Alwael, Saleh O. Bahaffi, Yousry M. Moustafa and Mohammad S. El-Shahawi Additional information is available at the end of the chapter

Dyab A. Al-Eryani, Waqas Ahmad, Zeinab M. Saigl,

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/intechopen.73397

#### **Abstract**

[45] Fedotov PS, Kördel W, Miró M, Peijnenburg WJGM, Wennrich R, Huang P-M. Extraction and fractionation methods for exposure assessment of trace metals, metalloids, and hazardous organic compounds in terrestrial environments. Critical Reviews in

[46] Hass A, Fine P. Sequential selective extraction procedures for the study of heavy metals in soils, sediments, and waste materials – A critical review. Critical Reviews in

[47] Krishnamurti GSR. Chemical methods for assessing contaminant bioavailability in soils. In: Naidu R, editor. Chemical Bioavailability in Terrestrial Environments. Oxford, UK:

[48] Kukina SE, Sadovnikova LK, Calafat-Frau A, Palerud R, Hummel H. Forms of metals in bottom sediments from some estuaries of the basins of the white and Barents seas.

[49] Tack FMG, Verloo MG. Chemical speciation and fractionation in soil and sediment heavy metal analysis: A review. International Journal of Environmental Analytical Chemistry.

[50] Ladonin DV. Heavy metal compounds in soils: Problems and methods of study. Eurasian

Environmental Science and Technology. 2012;**42**(11):1117-1171

Environmental Science and Technology. 2010;**40**(5):365-399

Geochemistry International. 1999;**37**(12):1197-1202

Elsevier; 2008. pp. 495-520

132 Trace Elements - Human Health and Environment

1995;**59**(2-4):225-238

Soil Science. 2002;**35**(6):605-613

An ultrasensitive, simple and highly selective spectrofluorometric strategy for quantifying traces of mercury(II) in environmental water has been established using the fluorescent probe procaine hydrochloride (PQ+ .Cl<sup>−</sup> ). The procedure was based upon the formation of the ternary ion associate complex [(PQ+ ) 2 .(HgI4 ) 2−] between PQ+ .Cl<sup>−</sup> and mercury(II) in iodide media at pH 9.0–10.0 with its subsequent extraction onto dichloromethane accompanied by a change in fluorescence intensity at λex/em = 268/333 nm. The developed strategy exhibited a linear range of 1–114 μg L−1 with lower limit of detection (LOD) and quantification (LOQ) of mercury(II) 1.3 and 3.98 nM, respectively. Intra and inter-day laboratory accuracy and precision for trace analysis of mercury(II) in water were performed. Complexed mercury(II) in environmental water, chemical speciation and successful literature comparison was performed. The proposed system offered excellent selectivity towards mercury(II) ions examined in the presence of competent ions in excess, relevant to real water samples. The method was applied for analysis of mercury(II) in tap water samples. Statistical comparison (Student's *t* and *F* tests) of the proposed method with the reference ICP-OES method revealed no significant differences in the accuracy and precision.

**Keywords:** spectrofluorometry, mercury(II), fluorescence, procaine hydrochloride, quenching, ternary ion associate complex

#### **1. Introduction**

Heavy metal pollution is global level environmental concern, which poses serious implications towards human health [1, 2]. Heavy metal pollution has attained considerable interest

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2018 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

in the recent past [3]. Among heavy metals, mercury is considered among the most toxic and dangerous ion due to its wide existence as an ore cinnabar in nature and its applications as pigment vermilion, detoxification/anticorrosive medicines and mercury fulminate detonator in explosives [3, 4]. Due to its wide presence in environment, it enters the biological membranes through respiratory and gastrointestinal tissues [4, 5]. It can also cause permanent harm to the endocrine and central nervous systems if accumulated in the food chain and ultimately in human body [5–13]. Moreover, a low-level exposure of mercury will affect the endocrine and nervous systems, brain and kidneys [14, 15].

Several forms of mercury including elemental, organic and inorganic or elemental display different levels of toxicity and contamination in natural water resources and drinking water [2]. The most stable and prevalent form of mercury that contributes to wider contamination is the solvated divalent mercuric ion (Hg2+), due to its high solubility in water [16]. The US Environmental Protection Agency (US-EPA) and World Health Organization (WHO) have set the permissible level (MPL) for mercury to 10 nM [1, 17]. Traces of mercury ions have shown significant toxicity, and therefore developing highly sensitive methodologies are considered essential [18]. Thus, recent trends have been oriented towards developing highly sensitive and selective procedures for monitoring and/or enrichment of mercury in various water samples [20–22].

**2. Experimental**

CH3

aqueous solution.

**2.2. Apparatus**

**2.1. Reagents and materials**

Glasswares were pre-cleaned with HNO3

**Figure 1.** Chemical structure of procaine hydrochloride.

mercury(II) was prepared from HgCl2

mercury(II) and pH measurements, respectively.

**2.3. Recommended procedures**

*2.3.1. Spectrophotometric procedure*

(20% m/v), acetone and deionized water. Analytical

http://dx.doi.org/10.5772/intechopen.73397

135

(BDH Chemicals, Poole, England) in ultra-pure water.

BO3 , H3 PO4 ,

reagent grade chemicals and solvents were used as received. Low-density polyethylene (LDPE) bottles, Nalgene were used for storage of the stock solutions. A stock solution (1.0 mg mL−1) of standard mercury(II) solution was prepared in ultra-pure water. Stock solution (1000 μM) of

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor…

More dilute solutions of mercury(II) ions were prepared by series dilution in deionized water. Potassium iodide solution (10% w/v) was prepared in ultra-pure water. Stock solutions of other cations were synthesized either from their chloride or nitrate salts in deionized water.

COOH and NaOH in deionized water as reported [65]. A stock solution (1000 μM) of procaine hydrochloride (Sigma–Aldrich) was prepared by dissolving the required weight in

Fluorescent measurements were recorded on a Perkin-Elmer LS55 spectrofluorometer, USA equipped with quartz cuvettes of path length 10 × 10 mm. UV-visible (190–1100 nm) spectra were recorded on a Perkin-Elmer spectrophotometer (Lambda 25, Shelton, CT, USA). For method validation, a Perkin-Elmer inductively coupled plasma-optical emission spectrometry (ICP-OES) (California, CT, USA) was utilized for mercury determination at the optimum operational parameters (**Table 1**). Deionized water was obtained from Milli-Q Plus system (Millipore, Bedford, MA, USA). A digital micropipette (Plus-Sed) and pH meter (inoLap pH/ ion level 2) were used for preparation of stock and more diluted solutions of reagent and

Appropriate volume (1.0 mL) of various mercury(II) concentrations (1.0 × 10−5 to 8 × 10−5 M) was transferred to a series of glass test tubes (10.0 mL) followed by addition of 100 μL KI (10% w/v). The solutions were made up to the mark with Britton-Robinson (B-R) buffer solution of pH 10 after

Britton-Robinson (B-R) (pH 2–11) buffers were prepared from BDH purchased H<sup>3</sup>

Several analytical procedures, including atomic absorption and emission spectrometry (AAS, AES) [19, 20], inductively coupled plasma-optical emission spectrometry (ICP-OES) [21], ion exchange chromatography [22], mechanical filtration [23], chemical precipitation [24], reverse osmosis, flotation [25, 26] and membrane separation [27], are reported for mercury determination. On the other hand, numerous liquid and solid phase extraction methods such as liquid-liquid extraction (LLE) [28], carbon nanotubes [29], graphene oxide [30–33] and polymers [34–36] have been reported for routine analysis of mercury(II). However, most of these methods require sophisticated equipment unavailable for use in developing countries with numerous other limitations like high cost, complications in their proper operation, selectivity and sensitivity [34, 35].

Recently, several molecular probe-based sensors using organic chromophores, quantum dots (QDs), small fluorescent organic molecules, proteins, antibodies and conjugated polymers coupled with several spectrometric and electrochemical techniques are reported for mercury(II) determination [36–53]. Some of these methods suffere from solubility issues, low stability, lower sensitivity and selectivity, complicated synthesis procedures and environmentally unfriendliness to monitor mercury(II) in biological and environmental samples. Hence, the work presented in this chapter is focused on: (i) developing a simple, highly selective and sensitive extractive spectrofluorometric LLE for trace determination of mercury(II) species in water samples using the ion pairing reagent 4-amino-N-(2-diethylaminoethyl) benzamide hydrochloride namely procaine hydrochloride and abbreviated as (PQ+ .Cl<sup>−</sup> ) (**Figure 1**) [54–59]; (ii) the utility of the proposed extractive LLE for trace determination of mercury ions in environmental water samples and finally (iii) validation and assignment of the most probable stoichiometry and chemical equilibria and fluorescence quenching mechanism of the produced ternary ion associate complex [(PQ+ )2 .(HgI4 )2−] of [HgI4 ]2− with the proposed ion pairing reagent [60–64].

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor… http://dx.doi.org/10.5772/intechopen.73397 135

**Figure 1.** Chemical structure of procaine hydrochloride.

## **2. Experimental**

in the recent past [3]. Among heavy metals, mercury is considered among the most toxic and dangerous ion due to its wide existence as an ore cinnabar in nature and its applications as pigment vermilion, detoxification/anticorrosive medicines and mercury fulminate detonator in explosives [3, 4]. Due to its wide presence in environment, it enters the biological membranes through respiratory and gastrointestinal tissues [4, 5]. It can also cause permanent harm to the endocrine and central nervous systems if accumulated in the food chain and ultimately in human body [5–13]. Moreover, a low-level exposure of mercury will affect the endocrine and

Several forms of mercury including elemental, organic and inorganic or elemental display different levels of toxicity and contamination in natural water resources and drinking water [2]. The most stable and prevalent form of mercury that contributes to wider contamination is the solvated divalent mercuric ion (Hg2+), due to its high solubility in water [16]. The US Environmental Protection Agency (US-EPA) and World Health Organization (WHO) have set the permissible level (MPL) for mercury to 10 nM [1, 17]. Traces of mercury ions have shown significant toxicity, and therefore developing highly sensitive methodologies are considered essential [18]. Thus, recent trends have been oriented towards developing highly sensitive and selective procedures for monitoring and/or enrichment of mercury in various water sam-

Several analytical procedures, including atomic absorption and emission spectrometry (AAS, AES) [19, 20], inductively coupled plasma-optical emission spectrometry (ICP-OES) [21], ion exchange chromatography [22], mechanical filtration [23], chemical precipitation [24], reverse osmosis, flotation [25, 26] and membrane separation [27], are reported for mercury determination. On the other hand, numerous liquid and solid phase extraction methods such as liquid-liquid extraction (LLE) [28], carbon nanotubes [29], graphene oxide [30–33] and polymers [34–36] have been reported for routine analysis of mercury(II). However, most of these methods require sophisticated equipment unavailable for use in developing countries with numerous other limitations like high cost, complications in their proper operation, selectivity

Recently, several molecular probe-based sensors using organic chromophores, quantum dots (QDs), small fluorescent organic molecules, proteins, antibodies and conjugated polymers coupled with several spectrometric and electrochemical techniques are reported for mercury(II) determination [36–53]. Some of these methods suffere from solubility issues, low stability, lower sensitivity and selectivity, complicated synthesis procedures and environmentally unfriendliness to monitor mercury(II) in biological and environmental samples. Hence, the work presented in this chapter is focused on: (i) developing a simple, highly selective and sensitive extractive spectrofluorometric LLE for trace determination of mercury(II) species in water samples using the ion pairing reagent 4-amino-N-(2-diethylaminoethyl) benzamide

(ii) the utility of the proposed extractive LLE for trace determination of mercury ions in environmental water samples and finally (iii) validation and assignment of the most probable stoichiometry and chemical equilibria and fluorescence quenching mechanism of the pro-

)2−] of [HgI4

)2 .(HgI4 .Cl<sup>−</sup>

]2− with the proposed ion pairing

) (**Figure 1**) [54–59];

hydrochloride namely procaine hydrochloride and abbreviated as (PQ+

duced ternary ion associate complex [(PQ+

nervous systems, brain and kidneys [14, 15].

134 Trace Elements - Human Health and Environment

ples [20–22].

and sensitivity [34, 35].

reagent [60–64].

#### **2.1. Reagents and materials**

Glasswares were pre-cleaned with HNO3 (20% m/v), acetone and deionized water. Analytical reagent grade chemicals and solvents were used as received. Low-density polyethylene (LDPE) bottles, Nalgene were used for storage of the stock solutions. A stock solution (1.0 mg mL−1) of standard mercury(II) solution was prepared in ultra-pure water. Stock solution (1000 μM) of mercury(II) was prepared from HgCl2 (BDH Chemicals, Poole, England) in ultra-pure water. More dilute solutions of mercury(II) ions were prepared by series dilution in deionized water. Potassium iodide solution (10% w/v) was prepared in ultra-pure water. Stock solutions of other cations were synthesized either from their chloride or nitrate salts in deionized water. Britton-Robinson (B-R) (pH 2–11) buffers were prepared from BDH purchased H<sup>3</sup> BO3 , H3 PO4 , CH3 COOH and NaOH in deionized water as reported [65]. A stock solution (1000 μM) of procaine hydrochloride (Sigma–Aldrich) was prepared by dissolving the required weight in aqueous solution.

#### **2.2. Apparatus**

Fluorescent measurements were recorded on a Perkin-Elmer LS55 spectrofluorometer, USA equipped with quartz cuvettes of path length 10 × 10 mm. UV-visible (190–1100 nm) spectra were recorded on a Perkin-Elmer spectrophotometer (Lambda 25, Shelton, CT, USA). For method validation, a Perkin-Elmer inductively coupled plasma-optical emission spectrometry (ICP-OES) (California, CT, USA) was utilized for mercury determination at the optimum operational parameters (**Table 1**). Deionized water was obtained from Milli-Q Plus system (Millipore, Bedford, MA, USA). A digital micropipette (Plus-Sed) and pH meter (inoLap pH/ ion level 2) were used for preparation of stock and more diluted solutions of reagent and mercury(II) and pH measurements, respectively.

#### **2.3. Recommended procedures**

#### *2.3.1. Spectrophotometric procedure*

Appropriate volume (1.0 mL) of various mercury(II) concentrations (1.0 × 10−5 to 8 × 10−5 M) was transferred to a series of glass test tubes (10.0 mL) followed by addition of 100 μL KI (10% w/v). The solutions were made up to the mark with Britton-Robinson (B-R) buffer solution of pH 10 after


**Table 1.** Operational parameters of ICP-OES for analysis of mercury.

adding 1.0 mL (1.0 × 10−4 M) of procaine hydrochloride. The solution mixtures were shaken well for 3.0 min with dichloromethane (5.0 mL). After equilibrium, the organic phase was separated out, shaken with anhydrous Na2 SO4 to remove trace of water and finally the absorbance of the organic extract was finally measured at λmax 310 nm *versus* the reagent blank at room temperature.

#### *2.3.2. Spectrofluorometric procedure*

In a series of glass test tubes (10.0 mL), appropriate concentrations (20–140 nM) of mercury(II) and 100 μL of 10% KI (w/v) were added. The test solutions were completed to the mark with B-R buffer of pH 10.0 after adding 1.0 mL (20 μM) of PQ<sup>+</sup> .Cl<sup>−</sup> and shaken well for 3.0 min with dichloromethane (5.0 mL). The organic phase was separated after equilibrium, shaken with anhydrous Na2 SO4 and analyzed at λex/em = 268/333 nm *versus* a reagent blank at room temperature. The quenched fluorescence in signal intensity (∆F) of PQ<sup>+</sup> .Cl<sup>−</sup> by mercury(II) was computed employing the following equation:

$$
\Delta \mathbf{F} = \mathbf{F}\_0 - \mathbf{F} \tag{1}
$$

*2.3.3. Calibration curve of mercury(II)*

**2.4. Analytical applications**

**3. Results and discussion**

associate complex [(PQ+

2 [PQ+

The strong fluorescence band of PQ<sup>+</sup>

**complex**

sured *versus* a reagent blank as described above.

added to water samples were finally computed.

The absorption spectrum of the ion pairing reagent PQ+

ate complex of mercury can be expressed as follows [68–73]:

]aq + [HgI4]

As shown in **Figure 3**, the solution of reagent blank PQ+

Hg2+ + 4I<sup>−</sup> ⇌ [HgI4]

)2−] of PQ+

2−

.Cl<sup>−</sup>

)2 .(HgI4

To construct the linear calibration plot, a series of solutions (10.0 mL) containing various

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor…

transferred to measuring flask (25 mL) prepared. The test solutions were completed to the mark with B-R buffer of pH 10 and shaken well with dichloromethane (5.0 mL) for 3.0 min.

analyzed at λex/em = 268/333 nm *versus* a reagent blank at room temperature. The quenched fluorescence signal intensity (∆F) of the formed colorless complex species was finally mea-

In LDPE sample bottles, pre-cleaned as described in Section 3.2.1, tap water samples were collected from the laboratories of the Department of Chemistry, King Abdul Aziz University, Jeddah, KSA and immediately filtered through 0.25 μm cellulose membrane filters before analysis and stored in LDPE bottles. The sample solutions were then spiked with known concentration (20–100 nM) of mercury(II). The fluorescence intensity of the test solutions were measured at (λex/em = 268/333 nm) under the optimized experimental conditions of mercury(II) by standard addition plot. The concentration and the percent recovery of the mercury(II)

**3.1. Electronic and fluorescence spectra of reagent and its mercury(1I) ion associate** 

.Cl<sup>−</sup>

The spectrum showed one well-defined band at 290 nm and was safely assigned to n → π\* electronic transitions [66, 67]. The electronic absorption spectra of the developed ternary ion

presence of an excess aqueous KI (10% w/v) were recorded and a well-defined absorption peak at 310 nm was noticed (**Figure 2**). The most probable formation mechanism of ion associ-

aq + nSorg ⇌ {[PQ+

exhibits a strong fluorescence intensity at λex/em = 268/333 nm in dichloromethane [68, 74–76].

increases after adding excess of mercury(II), through the formation of ternary ion associate

.Cl<sup>−</sup>

]2 .[HgI4] 2−

.Cl<sup>−</sup>

can be quenched after addition of 40 nM Hg(II) and

is demonstrated in **Figure 2**.

<sup>2</sup><sup>−</sup> (2)

.nS}org (3)

with concentration of 5.0 μM

(4 × 10−5 M) with mercury(II) (1 × 10−5 M) in the

After equilibrium, the organic phase was separated out, shaken with anhydrous Na2

.Cl<sup>−</sup>

http://dx.doi.org/10.5772/intechopen.73397

(1.0 mL, 20 μM) were

SO4 and 137

known (5.0 mL, 20–140 nM) concentrations of mercury(II) and PQ<sup>+</sup>

where F0 and F are the fluorescence intensities of the reagent before and after addition of mercury(II) under the optimized analytical parameters, respectively. The fluorescence signal intensity (F) of the formed ion associate complex was measured *versus* a reagent blank as described above. The quenched fluorescence (∆F) was finally determined. Following the recommended procedures, the selectivity of the proposed method was examined in the presence of a series of the concurrent diverse ions, e.g. Ca2+, Ba2+, Cu2+, Pb2+, Fe3+, As3+, Ag+ , Al3+ Sn2+, Cd2+, Bi3+, WO4 2−, MnO<sup>4</sup> − , F<sup>−</sup> , CO3 2−, SO4 2− in the presence of mercury(II) at 2.0 μg L−1.

#### *2.3.3. Calibration curve of mercury(II)*

To construct the linear calibration plot, a series of solutions (10.0 mL) containing various known (5.0 mL, 20–140 nM) concentrations of mercury(II) and PQ<sup>+</sup> .Cl<sup>−</sup> (1.0 mL, 20 μM) were transferred to measuring flask (25 mL) prepared. The test solutions were completed to the mark with B-R buffer of pH 10 and shaken well with dichloromethane (5.0 mL) for 3.0 min. After equilibrium, the organic phase was separated out, shaken with anhydrous Na2 SO4 and analyzed at λex/em = 268/333 nm *versus* a reagent blank at room temperature. The quenched fluorescence signal intensity (∆F) of the formed colorless complex species was finally measured *versus* a reagent blank as described above.

#### **2.4. Analytical applications**

In LDPE sample bottles, pre-cleaned as described in Section 3.2.1, tap water samples were collected from the laboratories of the Department of Chemistry, King Abdul Aziz University, Jeddah, KSA and immediately filtered through 0.25 μm cellulose membrane filters before analysis and stored in LDPE bottles. The sample solutions were then spiked with known concentration (20–100 nM) of mercury(II). The fluorescence intensity of the test solutions were measured at (λex/em = 268/333 nm) under the optimized experimental conditions of mercury(II) by standard addition plot. The concentration and the percent recovery of the mercury(II) added to water samples were finally computed.

## **3. Results and discussion**

adding 1.0 mL (1.0 × 10−4 M) of procaine hydrochloride. The solution mixtures were shaken well for 3.0 min with dichloromethane (5.0 mL). After equilibrium, the organic phase was separated

organic extract was finally measured at λmax 310 nm *versus* the reagent blank at room temperature.

In a series of glass test tubes (10.0 mL), appropriate concentrations (20–140 nM) of mercury(II) and 100 μL of 10% KI (w/v) were added. The test solutions were completed to the mark with

with dichloromethane (5.0 mL). The organic phase was separated after equilibrium, shaken

∆F = F0 –F (1)

mercury(II) under the optimized analytical parameters, respectively. The fluorescence signal intensity (F) of the formed ion associate complex was measured *versus* a reagent blank as described above. The quenched fluorescence (∆F) was finally determined. Following the recommended procedures, the selectivity of the proposed method was examined in the presence

of a series of the concurrent diverse ions, e.g. Ca2+, Ba2+, Cu2+, Pb2+, Fe3+, As3+, Ag+

and F are the fluorescence intensities of the reagent before and after addition of

2− in the presence of mercury(II) at 2.0 μg L−1.

to remove trace of water and finally the absorbance of the

.Cl<sup>−</sup>

and analyzed at λex/em = 268/333 nm *versus* a reagent blank at room

and shaken well for 3.0 min

by mercury(II) was

, Al3+ Sn2+, Cd2+,

.Cl<sup>−</sup>

SO4

Nebulizer type Cross-flow (Gim Tip)

Spray chamber Scott (Ryton) Injector Scott (Ryton) Analytical wavelength Hg 194.2 nm

**Table 1.** Operational parameters of ICP-OES for analysis of mercury.

**Parameter Unit** RF power 1400 W Nebulizer flow 0.7 L/min Auxiliary flow 0.3 L/min Plasma flow 10.0 L/min Sample pump flow 1 mL/min Plasma viewing Axial Processing mode Area Replicates 3

136 Trace Elements - Human Health and Environment

B-R buffer of pH 10.0 after adding 1.0 mL (20 μM) of PQ<sup>+</sup>

2−, SO4

temperature. The quenched fluorescence in signal intensity (∆F) of PQ<sup>+</sup>

SO4

computed employing the following equation:

out, shaken with anhydrous Na2

*2.3.2. Spectrofluorometric procedure*

with anhydrous Na2

where F0

Bi3+, WO4

2−, MnO<sup>4</sup> − , F<sup>−</sup> , CO3

#### **3.1. Electronic and fluorescence spectra of reagent and its mercury(1I) ion associate complex**

The absorption spectrum of the ion pairing reagent PQ+ .Cl<sup>−</sup> is demonstrated in **Figure 2**. The spectrum showed one well-defined band at 290 nm and was safely assigned to n → π\* electronic transitions [66, 67]. The electronic absorption spectra of the developed ternary ion associate complex [(PQ+ )2 .(HgI4 ) 2−] of PQ+ .Cl<sup>−</sup> (4 × 10−5 M) with mercury(II) (1 × 10−5 M) in the presence of an excess aqueous KI (10% w/v) were recorded and a well-defined absorption peak at 310 nm was noticed (**Figure 2**). The most probable formation mechanism of ion associate complex of mercury can be expressed as follows [68–73]:

$$\mathrm{Hg}^{2+} + 4\mathrm{I}^{-} \rightleftharpoons \ [\mathrm{HgI}\_{4}]^{2-} \tag{2}$$

$$2\left[\mathrm{PQ}^{\circ}\right]\_{\mathrm{aq}} + \left[\mathrm{H}\mathrm{gI}\_{4}\right]^{2-}\_{\mathrm{aq}} + \mathrm{nS}\_{\mathrm{org}} \rightleftharpoons \left\{\left[\mathrm{PQ}^{\circ}\right]\_{2}\left[\mathrm{H}\mathrm{gI}\_{4}\right]^{2-}\mathrm{nS}\right\}\_{\mathrm{aq}}\tag{3}$$

As shown in **Figure 3**, the solution of reagent blank PQ+ .Cl<sup>−</sup> with concentration of 5.0 μM exhibits a strong fluorescence intensity at λex/em = 268/333 nm in dichloromethane [68, 74–76]. The strong fluorescence band of PQ<sup>+</sup> .Cl<sup>−</sup> can be quenched after addition of 40 nM Hg(II) and increases after adding excess of mercury(II), through the formation of ternary ion associate

**Figure 2.** Electronic spectra of the PQ+ .Cl<sup>−</sup> (4 × 10−5 M) (1) and its ternary ion associate complex [(PQ<sup>+</sup> ) 2 .(HgI4 )2−] in dichloromethane (2). Condition: Mercury (1 × 10−5 M).

quenching was observed in dichloromethane. Therefore, a detailed study on the influence of

was critically studied (**Figure 5**). Thus, dichloromethane was adopted at 2.0 mL volume in the subsequent work pertaining to its higher performance compared to other volume fractions.

studied in B-R (pH 3–12) buffer solutions. Maximum quenching by mercury(II) quencher was achieved at pH ≈ 9–10. The extraction rate at pH < 10.0 for ternary ion associate complex

amount of mercury extracted at pH values other than pH ≈ 9–10 [62, 63]. Thus, in the next work, the solution media were adopted at pH ≈ 10 due to the ease in formation of the ternary

)2−] was lower. Several factors including hydrolysis, instability and/or incomplete

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor…

[HgI4

The effect of pH on the fluorescence quenching of the formed mercury(II)-PQ<sup>+</sup>

.Cl<sup>−</sup>

.Cl<sup>−</sup>

(5.0 μM) reagent by mercury

(5.0 μM) reagent by K<sup>2</sup>

http://dx.doi.org/10.5772/intechopen.73397

.Cl<sup>−</sup>

reagent by mercury(II)

[HgI4

] (80 nM).

139

]2− account for the decrease in the

complex was

dichloromethane volume on the fluorescence quenching of PQ<sup>+</sup>

**Figure 4.** Influence of type extraction solvent on the fluorescence quenching of PQ<sup>+</sup>

extraction, and the slow dissociation of the complex K2

) 2 .(HgI4 )2−].

**Figure 5.** Influence of dichloromethane volume on the fluorescence quenching of PQ<sup>+</sup>

*3.2.2. Effect of pH*

ion associate complex [(PQ+

[(PQ+ )2 .(HgI4

(80 nM).

**Figure 3.** Fluorescence spectra of PQ+ .Cl<sup>−</sup> (5.0 μM) and its ternary ion associate complex [(PQ<sup>+</sup> ) 2 .(HgI4 )2−] with mercury(II) ions (40, 80 nM).

complex [(PQ+ )2 .(HgI4 )2−] (**Figure 3**) between PQ+ .Cl<sup>−</sup> and HgI4 2− [60]. Hence, this sensing principle by virtue of quenching was successfully applied for sensitive spectrofluorometric determination of mercury(II).

#### **3.2. Analytical parameters**

The effect of various experimental conditions on the fluorescence characteristics of both ion associate complex was studied, including type of solvents, pH, reagent concentration and equilibrium time.

#### *3.2.1. Effect of solvent*

The influence of extraction solvent, e.g. n-hexane, chloroform, dichloromethane, toluene and cyclohexane, on the fluorescence quenching of the formed mercury(II) complex was studied. The results are demonstrated in **Figure 4**. Stable and maximum change in fluorescence

**Figure 4.** Influence of type extraction solvent on the fluorescence quenching of PQ<sup>+</sup> .Cl<sup>−</sup> (5.0 μM) reagent by K<sup>2</sup> [HgI4 ] (80 nM).

quenching was observed in dichloromethane. Therefore, a detailed study on the influence of dichloromethane volume on the fluorescence quenching of PQ<sup>+</sup> .Cl<sup>−</sup> reagent by mercury(II) was critically studied (**Figure 5**). Thus, dichloromethane was adopted at 2.0 mL volume in the subsequent work pertaining to its higher performance compared to other volume fractions.

#### *3.2.2. Effect of pH*

complex [(PQ+

ions (40, 80 nM).

equilibrium time.

*3.2.1. Effect of solvent*

)2 .(HgI4

**Figure 3.** Fluorescence spectra of PQ+

**Figure 2.** Electronic spectra of the PQ+

138 Trace Elements - Human Health and Environment

dichloromethane (2). Condition: Mercury (1 × 10−5 M).

mination of mercury(II).

**3.2. Analytical parameters**

)2−] (**Figure 3**) between PQ+

.Cl<sup>−</sup>

.Cl<sup>−</sup>

.Cl<sup>−</sup>

(5.0 μM) and its ternary ion associate complex [(PQ<sup>+</sup>

ciple by virtue of quenching was successfully applied for sensitive spectrofluorometric deter-

The effect of various experimental conditions on the fluorescence characteristics of both ion associate complex was studied, including type of solvents, pH, reagent concentration and

The influence of extraction solvent, e.g. n-hexane, chloroform, dichloromethane, toluene and cyclohexane, on the fluorescence quenching of the formed mercury(II) complex was studied. The results are demonstrated in **Figure 4**. Stable and maximum change in fluorescence

and HgI4

(4 × 10−5 M) (1) and its ternary ion associate complex [(PQ<sup>+</sup>

2− [60]. Hence, this sensing prin-

2−] with mercury(II)

) 2 .(HgI4 ) ) 2 .(HgI4 )2−] in

> The effect of pH on the fluorescence quenching of the formed mercury(II)-PQ<sup>+</sup> complex was studied in B-R (pH 3–12) buffer solutions. Maximum quenching by mercury(II) quencher was achieved at pH ≈ 9–10. The extraction rate at pH < 10.0 for ternary ion associate complex [(PQ+ ) 2 .(HgI4 )2−] was lower. Several factors including hydrolysis, instability and/or incomplete extraction, and the slow dissociation of the complex K2 [HgI4 ]2− account for the decrease in the amount of mercury extracted at pH values other than pH ≈ 9–10 [62, 63]. Thus, in the next work, the solution media were adopted at pH ≈ 10 due to the ease in formation of the ternary ion associate complex [(PQ+ )2 .(HgI4 )2−].

**Figure 5.** Influence of dichloromethane volume on the fluorescence quenching of PQ<sup>+</sup> .Cl<sup>−</sup> (5.0 μM) reagent by mercury (80 nM).

#### *3.2.3. Effect of extraction time*

The stability and the fluorescence signal intensity of the emission spectrum of the formed ternary ion associate complex [(PQ+ ) 2 .(HgI4 )2−] considerably depend on the reaction time. Therefore, the fluorescence intensity of the emission spectrum of the ternary ion associate complex [(PQ+ )2 .(HgI4 )2−] was measured at various time intervals (0.5–14 min) at λex/em = 268/333 nm at the optimum conditions. The results are demonstrated in **Figure 6**. Maximum stability in signal intensity was achieved after 2.0 min and remained constant for longer time up to 14.0 min. Therefore, a standing time of 2.0 min was adopted in the subsequent work.

#### *3.2.4. Effect of PQ<sup>+</sup> .Cl<sup>−</sup> concentration*

The influence of PQ<sup>+</sup> .Cl<sup>−</sup> concentration on the fluorescence quenching of mercury(II) at concentration of 80 nM was studied in KI (0.1% w/v). Thus, various fractions of PQ<sup>+</sup> .Cl<sup>−</sup> solution were added to [HgI4 ]2−. The fluorescence quenching (∆F) increased on increasing PQ<sup>+</sup> .Cl<sup>−</sup> concentration up to 5.0 μM and leveled off at higher concentration. This possible self-absorption and aggregation of the reagent at high concentration contribute to this enhanced behavior [77]. Thus, in the subsequent work, a concentration of 5.0 μM of PQ<sup>+</sup> .Cl<sup>−</sup> was selected, capable of successfully quantifying the target up to trace level proportions.

PQ+

interfering ions in higher concentration.

**Figure 7.** The values of quenching fluorescence intensity of PQ<sup>+</sup>

presence of different metal ions (500 μgL−1) in KI (0.1% w/v).

**3.4. Analytical performance**

lowing regression equation:

respectively.

.Cl<sup>−</sup> possess good selectivity in probing mercury(II). As shown in **Figure 7**, the signal intensity in the presence of 2.0 μg L−1 mercury(II) mixed with 500 μg L−1 of other metal ions in KI (0.1% w/v) was found similar to mercury(II) alone at 2.0 μg L−1 in KI (0.1% w/v). Thus, it can be concluded that the probe is selective towards mercury(II) in the presence of potent

.Cl<sup>−</sup>

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor…

Under the optimized experimental conditions, the influence of various known mercury(II)

plot of mercury(II) concentrations (20–140 nM) *versus* ∆F was linear (**Figure 8**) with the fol-

∆F = 1.6354 [C] (nM)–3.3971 (4)

where C and ∆F are the concentration of mercury(II) ions and quenched fluorescence,

The precision of the method was checked by means of ANOVA by performing five successive replicates per day for 5 days for a typical sample containing 100 nM mercury(II) (**Table 2**).

The precision of the proposed method was evaluated. The calculated value of *F* (F = 1.64) (**Table 3**) was found lower than the tabulated *F* value (F = 2.87) at 95% confidence level [78]. The obtained LOD and LOQ [78] for the proposed method for mercury(II) species were 1.3 and 3.98 nM, respectively. This value was significantly lower than the maximum allowable mercury concentration (10.0 nM) by USEPA in drinking water [17]. The proposed method

was recorded. The

(5.0 μgL−1) with addition of Hg(II) (2.0 μgL−1) in the

http://dx.doi.org/10.5772/intechopen.73397

141

concentrations (20–140 nM) on the fluorescence spectra of mercury(II)-PQ<sup>+</sup>

#### **3.3. Selectivity**

Microenvironment resembling real sample matrix containing competent interfering ion was designed to critically examine the applicability of the proposed method. Thus, the quenching in the fluorescence intensity of the ternary ion associate complex [(PQ<sup>+</sup> )2 .(HgI4 )2−] before and after adding the interfering ions in the presence of relatively high concentration (50–500 μgL−1) of other metal ions (K+ , Ca2+, Ba2+, Cu2+, Pb2+, Fe3+, As3+, Ag+ , Al3+ Sn2+, Cd2+, Bi3+, WO4 2−, MnO<sup>4</sup> − , F<sup>−</sup> , CO3 2−, SO4 2−) under the same condition was studied. The results are shown in **Figure 7**, where no noticeable quenching changes were observed. The data indicated that

**Figure 6.** Plot of time *vs.* the fluorescence quenching of PQ<sup>+</sup> .Cl<sup>−</sup> (5.0 μM) reagent by mercury (80 nM).

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor… http://dx.doi.org/10.5772/intechopen.73397 141

**Figure 7.** The values of quenching fluorescence intensity of PQ<sup>+</sup> .Cl<sup>−</sup> (5.0 μgL−1) with addition of Hg(II) (2.0 μgL−1) in the presence of different metal ions (500 μgL−1) in KI (0.1% w/v).

PQ+ .Cl<sup>−</sup> possess good selectivity in probing mercury(II). As shown in **Figure 7**, the signal intensity in the presence of 2.0 μg L−1 mercury(II) mixed with 500 μg L−1 of other metal ions in KI (0.1% w/v) was found similar to mercury(II) alone at 2.0 μg L−1 in KI (0.1% w/v). Thus, it can be concluded that the probe is selective towards mercury(II) in the presence of potent interfering ions in higher concentration.

#### **3.4. Analytical performance**

*3.2.3. Effect of extraction time*

140 Trace Elements - Human Health and Environment

plex [(PQ+

)2 .(HgI4

*3.2.4. Effect of PQ<sup>+</sup>*

The influence of PQ<sup>+</sup>

were added to [HgI4

**3.3. Selectivity**

MnO<sup>4</sup> − , F<sup>−</sup> , CO3

μgL−1) of other metal ions (K+

2−, SO4

**Figure 6.** Plot of time *vs.* the fluorescence quenching of PQ<sup>+</sup>

ternary ion associate complex [(PQ+

*.Cl<sup>−</sup>*

.Cl<sup>−</sup>

 *concentration*

The stability and the fluorescence signal intensity of the emission spectrum of the formed

Therefore, the fluorescence intensity of the emission spectrum of the ternary ion associate com-

at the optimum conditions. The results are demonstrated in **Figure 6**. Maximum stability in signal intensity was achieved after 2.0 min and remained constant for longer time up to

centration up to 5.0 μM and leveled off at higher concentration. This possible self-absorption and aggregation of the reagent at high concentration contribute to this enhanced behavior

Microenvironment resembling real sample matrix containing competent interfering ion was designed to critically examine the applicability of the proposed method. Thus, the quench-

and after adding the interfering ions in the presence of relatively high concentration (50–500

**Figure 7**, where no noticeable quenching changes were observed. The data indicated that

.Cl<sup>−</sup>

, Ca2+, Ba2+, Cu2+, Pb2+, Fe3+, As3+, Ag+

14.0 min. Therefore, a standing time of 2.0 min was adopted in the subsequent work.

centration of 80 nM was studied in KI (0.1% w/v). Thus, various fractions of PQ<sup>+</sup>

[77]. Thus, in the subsequent work, a concentration of 5.0 μM of PQ<sup>+</sup>

of successfully quantifying the target up to trace level proportions.

ing in the fluorescence intensity of the ternary ion associate complex [(PQ<sup>+</sup>

)2−] was measured at various time intervals (0.5–14 min) at λex/em = 268/333 nm

concentration on the fluorescence quenching of mercury(II) at con-

2−) under the same condition was studied. The results are shown in

(5.0 μM) reagent by mercury (80 nM).

]2−. The fluorescence quenching (∆F) increased on increasing PQ<sup>+</sup>

)2−] considerably depend on the reaction time.

.Cl<sup>−</sup>

.Cl<sup>−</sup>

was selected, capable

)2 .(HgI4

, Al3+ Sn2+, Cd2+, Bi3+, WO4

solution

)2−] before

2−,

.Cl<sup>−</sup> con-

) 2 .(HgI4

> Under the optimized experimental conditions, the influence of various known mercury(II) concentrations (20–140 nM) on the fluorescence spectra of mercury(II)-PQ<sup>+</sup> was recorded. The plot of mercury(II) concentrations (20–140 nM) *versus* ∆F was linear (**Figure 8**) with the following regression equation:

$$
\Delta \mathbf{F} = 1.6354 \,\mathrm{[C]} \,\mathrm{(nM)} \,\mathrm{-3.3971} \tag{4}
$$

where C and ∆F are the concentration of mercury(II) ions and quenched fluorescence, respectively.

The precision of the method was checked by means of ANOVA by performing five successive replicates per day for 5 days for a typical sample containing 100 nM mercury(II) (**Table 2**).

The precision of the proposed method was evaluated. The calculated value of *F* (F = 1.64) (**Table 3**) was found lower than the tabulated *F* value (F = 2.87) at 95% confidence level [78]. The obtained LOD and LOQ [78] for the proposed method for mercury(II) species were 1.3 and 3.98 nM, respectively. This value was significantly lower than the maximum allowable mercury concentration (10.0 nM) by USEPA in drinking water [17]. The proposed method

**Source of variation Sum of squares (SS) Degrees of freedom (df) Mean square (MS) F value S1**

.

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor…

Between days 41.55 4 10.39 1.64

**Technique Reagent LOD Ref** Fluorescence (2-Pyridylmethyl)(2-quinolylmethyl) amine 2.6 × 10−8 M [37]

Fluorescence Phenylamine-oligothiophene 4.39 × 10−7 M [39] Fluorescence Squaraine–bis(rhodamine-B) 6.47 × 10−6 M [40] Fluorescence Rhodamine (R-2) 1 × 10−8 M [41] Electrochemical sensor DNA-generated gold amalgam 0.002 × 10−9 M [42] Fluorescence Rhodamine 6G derivative and AuNPs 0.75 × 10−9 M [43] Fluorescence 1,4-Bis(styryl)benzene 7 × 10−9 M [44] Colorimetry Au-NPs 0.0084 × 10−9 M [45] Chemiluminescence Rhodamine B and gold nanoprisms 0.027 × 10−9 M [46]

2-hydroxy-acetophenone

Within days 126.77 20 6.34

: Mean square error.

**Table 3.** Analysis of variance (ANOVA) for the linear equation results†

Fluorescence Rhodamine hydrazine and

Single-crystal X-ray diffraction 1-[(5-Benzyl-1,3-thiazol-2-yl) diazenyl]

Colorimetry Carrageenan-functionalized Ag/AgCl

Electrochemiluminescence biosensor tris-(Bipyridine) (Ru(bpy) 3

Differential pulse voltammetry Polypyrrole decorated

Inductively coupled plasma-optical emission spectrometry (ICP-OES)

Electrochemical and X-ray photoelectron spectroscopy(XPS) naphthalene-2-ol

nanoparticles

Nafion

Spectrophotometric bis(2-Ethylhexyl) phosphate 3.5 × 10−9 M [48]

X-ray fluorescence (XRF) — 37 × 10−6 M [51]

Differential pulse voltammetry Copper cobalt hexacyanoferrate 80 × 10−9 M [52]

Fluorescence DNA-functionalized-graphene 4.1 × 10−9 M [30]

**Table 4.** Analytical features of different methods employed for mercury(II) determination.

Spectrofluorometry Procaine hydrochloride 1.3 × 10−9 M Present work

graphene/β-cyclodextrin

2+)/

— 0.15 × 10−9 M [21]

1-Undecanethiol assembled Au substrate 4.5 × 10−9 M [53]

cyclodextrins-Au nanoparticles(CD-AuNps)/

Total 168.32 24

: Regression mean square. **S2**

† **S1** **/S2**

143

150 × 10−9 M [38]

http://dx.doi.org/10.5772/intechopen.73397

0.41 × 10−6 M [47]

1 × 10−6 M [49]

0.1 × 10−9 M [50]

0.47 × 10−9 M [36]

**Figure 8.** Calibration plot of ternary ion associate complex [(PQ+ ) 2 .(HgI4 )2−]. Conditions: Mercury(II) (20–140 nM), PQ<sup>+</sup> (5.0 μM) and KI (0.1% w/v) pH ≈ 10 at λex/em = 268/333 nm.

could be considered for routine analysis of mercury due to high precision and selectivity in real samples in the presence of excess relevant competent ions. Moreover, the utility of the proposed method was finally evaluated by comparing the analytical features of the proposed method with a wide range of promising studies in literature. It includes successful comparison with a series of published fluorescence [30, 37–41, 43], electrochemical sensor [42], spectrophotometric [44, 48], colorimetry [45, 49], chemiluminescence [46], single-crystal X-ray diffraction [47], electrochemiluminescence biosensor [50], X-ray fluorescence (XRF) [51], differential pulse voltammetry [36, 52], ICP-OES [21], electrochemical [50] and X-ray photoelectron spectroscopy (XPS) [53], in terms of LOD and LOQ (**Table 4**).

#### **3.5. Fluorescence quenching mechanism**

The fluorescence quenching process of PQ<sup>+</sup> .Cl<sup>−</sup> by mercury(II) as a quencher was critically investigated to evaluate the fluorescence mechanism for the formed ternary ion associate complex [(PQ+ )2 .(HgI4 )2−] [60, 62–64]. The fluorescence spectra of PQ<sup>+</sup> .Cl<sup>−</sup> (λex/λem 268/330 nm)


**Table 2.** Five days and five replicates per day determined the quenching fluorescence intensity of the reagent PQ.Cl− by mercury(II) (100 nM) added.


**Table 3.** Analysis of variance (ANOVA) for the linear equation results† .

could be considered for routine analysis of mercury due to high precision and selectivity in real samples in the presence of excess relevant competent ions. Moreover, the utility of the proposed method was finally evaluated by comparing the analytical features of the proposed method with a wide range of promising studies in literature. It includes successful comparison with a series of published fluorescence [30, 37–41, 43], electrochemical sensor [42], spectrophotometric [44, 48], colorimetry [45, 49], chemiluminescence [46], single-crystal X-ray diffraction [47], electrochemiluminescence biosensor [50], X-ray fluorescence (XRF) [51], differential pulse voltammetry [36, 52], ICP-OES [21], electrochemical [50] and X-ray

) 2 .(HgI4

.Cl<sup>−</sup>

)2−] [60, 62–64]. The fluorescence spectra of PQ<sup>+</sup>

**Table 2.** Five days and five replicates per day determined the quenching fluorescence intensity of the reagent PQ.Cl−

**Replicate First day ∆F Second day ∆F Third day ∆F Fourth day ∆F Fifth day ∆F** 167.00 163.88 166.10 159.40 167.59 164.47 168.70 163.75 162.32 165.55 161.81 166.43 167.99 166.78 166.34 160.53 166.45 161.69 164.76 164.00 165.70 162.44 165.08 161.13 169.45 **Mean** 163.90 165.58 164.92 162.88 166.59 **SD** 2.69 2.45 2.38 2.93 2.06

investigated to evaluate the fluorescence mechanism for the formed ternary ion associate

by mercury(II) as a quencher was critically

)2−]. Conditions: Mercury(II) (20–140 nM), PQ<sup>+</sup>

.Cl<sup>−</sup>

(λex/λem 268/330 nm)

by

photoelectron spectroscopy (XPS) [53], in terms of LOD and LOQ (**Table 4**).

**3.5. Fluorescence quenching mechanism**

)2 .(HgI4

complex [(PQ+

mercury(II) (100 nM) added.

The fluorescence quenching process of PQ<sup>+</sup>

**Figure 8.** Calibration plot of ternary ion associate complex [(PQ+

(5.0 μM) and KI (0.1% w/v) pH ≈ 10 at λex/em = 268/333 nm.

142 Trace Elements - Human Health and Environment


**Table 4.** Analytical features of different methods employed for mercury(II) determination.

upon introduction of varying concentrations (0.3–1.0 μg L−1) of quencher (mercury(II) ions) are shown in **Figure 9**. The fluorescence intensity of PQ<sup>+</sup> .Cl<sup>−</sup> decreases regularly with increasing quencher concentration. The Stern-Volmer (KSV) constant was calculated by employing the equation:

$$\mathbf{F}\_0/\mathbf{F} = \mathbf{1} + \mathbf{K}\_\text{SV} \begin{bmatrix} \mathbf{Q} \end{bmatrix} \tag{5}$$

where F0 and F are the fluorescence signals in the absence and presence of [HgI<sup>4</sup> ]2− quencher, respectively. KSV is the Stern-Volmer constant and [Q] is the quencher concentration. The values of KSV and correlation factor calculated by plotting fluorescence quenching (ΔF) of PQ<sup>+</sup> . Cl<sup>−</sup> *versus* [Hg2+] were found equal to 1.87 × 10<sup>6</sup> L g−1 mol−1 and 0.9909, respectively.

The chemical composition of the ternary ion associate complex [(PQ+ )2 .(HgI4 )2−] was determined from the Benesi-Hildebrand linear model by employing the following equation [77, 79–82]:

[11] : \Delta\text{-}8\text{I}\text{'}: \tag{14.7-87}
$$\frac{1}{\{F\_s - F\}} = \frac{1}{\{F - F\_\perp\}} + \frac{1}{\{F - F\_\perp\} \times K \times [Q]^2} \tag{6}$$

**3.6. Analytical applications**

Average ± standard deviation (n = 5).

†

**Samples Spiked** 

**(nM/L)**

**4. Conclusion**

the fluorescence probe PQ<sup>+</sup>

the fluorescence quenching of PQ<sup>+</sup>

Applications of the proposed method were tested for analysis of mercury(II) in tap water samples (King Abdulaziz University, Jeddah, KSA). Samples were spiked with known concentrations (20–100 nM) of mercury(II) ion and analyzed by the developed method. In each sample,

**Proposed method Proposed PQ+** 

Tap water 1 20 20.97 ± 0.50 104.7 20.54 ± 0.63 102.7 Tap water 2 40 39.11 ± 1.25 97.78 39.22 ± 0.92 98.05 Tap water 3 60 61.74 ± 3.07 102.23 62.58 ± 0.72 104.3 Tap water 4 80 78.51 ± 1.63 98.14 77.88 ± 1.21 97.35 Tap water 5 100 97.74 ± 3.51 97.74 98.37 ± 1.30 98.37

**Detected (nM) mean ±** *SD*

**method**

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor…

onto the water samples. For method validation, mercury(II) concentrations in the spiked samples were also determined by the standard ICP-OES method. The results are summarized in **Table 5**. The recovery percentage of the measured mercury(II) added to the real samples by the developed and the standard ICP-OES methods was in the range from 97.74 to 104.7% and 98.05 to 102.7%, respectively. The calculated values of the Student's *t* and *F* tests were found lower than the tabulated Student's *t* and *F* tests at 0.05 probability [78] revealing no significant differences between both methods. Thus, it can be concluded that the proposed fluorescence

probe can be used as a potential assay for sensing of mercury(II) in complex matrices.

A new and facile extractive spectrofluorometric method for cost effective, precise, accurate and selective determination of trace levels of mercury(II) in water. The proposed method was

mercury(II) ions over most anions and metal cations. The detection process could be performed quickly at room temperature without any catalyst or oxidizer. The proposed method provides LOD (1.3 nM) lower than the value set by WHO (10 nM) and USEPA for drinkable water [1, 17]. The developed method is easy to operate as it does not require sophisticated experimental techniques, and the proposed assay is useful for point-of care applications. Moreover, the method opens capable ways for developing fluorescence assay strategies. The proposed approach was validated successfully by analysis of mercury(II) in environmental

was immediately measured after spiking of mercury(II)

.

**ICP-OES**

http://dx.doi.org/10.5772/intechopen.73397

**mean ±** *SD*

**Recovery (%)**

145

**Recovery (%) Detected (nM)** 

)2 .(HgI4 )

. The proposed system offered excellent selectivity towards

2−] between [HgI4

] 2− and

.Cl<sup>−</sup>

**Table 5.** Analytical data mercury(II) assay by the developed and ICP-OES methods†

based upon formation of ternary ion associate complex [(PQ+

.Cl<sup>−</sup>

where *F∞* represents the emission intensity of the ternary ion associate complex [(PQ+ )2 . (HgI4 )2−] at equilibrium and *K* is association constant. The number of the binding sites (n) and the apparent binding constant (K) of PQ+ .Cl<sup>−</sup> that independently binds to equivalent sites on a macromolecule were determined from the linear plot of Benesi-Hildebrand (1/(I-Iₒ) *versus* 1/ ([Mercury(II)]). The plot revealed formation of 1:2 stoichiometry of [HgI<sup>4</sup> ]:PQ+ .Cl<sup>−</sup> molar ratio in the produced ternary ion associate complex [(PQ+ ) 2 .(HgI4 )2−]. The calculated association constant *K* was found equal to 73 M−1.

**Figure 9.** Fluorescence quenching spectra of PQ+ .Cl<sup>−</sup> (5.0 μM) in the presence of various concentrations (20–80 nM) of the mercury(II) ions. Conditions: KI (0.1% w/v) pH ≈ 10, at λex/em = 268/333 nm.

A Simple and Highly Structured Procaine Hydrochloride as Fluorescent Quenching Chemosensor… http://dx.doi.org/10.5772/intechopen.73397 145


**Table 5.** Analytical data mercury(II) assay by the developed and ICP-OES methods† .

#### **3.6. Analytical applications**

upon introduction of varying concentrations (0.3–1.0 μg L−1) of quencher (mercury(II) ions)

ing quencher concentration. The Stern-Volmer (KSV) constant was calculated by employing

F0 /F = 1 + KSV [Q] (5)

respectively. KSV is the Stern-Volmer constant and [Q] is the quencher concentration. The values of KSV and correlation factor calculated by plotting fluorescence quenching (ΔF) of PQ<sup>+</sup>

mined from the Benesi-Hildebrand linear model by employing the following equation

where *F∞* represents the emission intensity of the ternary ion associate complex [(PQ+

a macromolecule were determined from the linear plot of Benesi-Hildebrand (1/(I-Iₒ) *versus* 1/

.Cl<sup>−</sup>

)2−] at equilibrium and *K* is association constant. The number of the binding sites (n) and

) 2 .(HgI4 )

and F are the fluorescence signals in the absence and presence of [HgI<sup>4</sup>

Cl<sup>−</sup> *versus* [Hg2+] were found equal to 1.87 × 10<sup>6</sup> L g−1 mol−1 and 0.9909, respectively.

The chemical composition of the ternary ion associate complex [(PQ+

(*Fo* <sup>−</sup> *<sup>F</sup>*) <sup>=</sup> \_\_\_\_\_\_ <sup>1</sup>

([Mercury(II)]). The plot revealed formation of 1:2 stoichiometry of [HgI<sup>4</sup>

.Cl<sup>−</sup>

mercury(II) ions. Conditions: KI (0.1% w/v) pH ≈ 10, at λex/em = 268/333 nm.

.Cl<sup>−</sup>

decreases regularly with increas-

)2 .(HgI4

(*<sup>F</sup>* <sup>−</sup> *<sup>F</sup>*∞) <sup>+</sup> \_\_\_\_\_\_\_\_\_\_\_\_\_ 1 (*<sup>F</sup>* <sup>−</sup> *<sup>F</sup>*∞) <sup>×</sup> *<sup>K</sup>* <sup>×</sup> [*Q*]2 (6)

that independently binds to equivalent sites on

(5.0 μM) in the presence of various concentrations (20–80 nM) of the

]:PQ+ .Cl<sup>−</sup>

2−]. The calculated association

]2− quencher,

)2−] was deter-

.

)2 .

molar ratio

are shown in **Figure 9**. The fluorescence intensity of PQ<sup>+</sup>

the equation:

where F0

[77, 79–82]:

(HgI4

\_\_\_\_\_ <sup>1</sup>

144 Trace Elements - Human Health and Environment

the apparent binding constant (K) of PQ+

constant *K* was found equal to 73 M−1.

**Figure 9.** Fluorescence quenching spectra of PQ+

in the produced ternary ion associate complex [(PQ+

Applications of the proposed method were tested for analysis of mercury(II) in tap water samples (King Abdulaziz University, Jeddah, KSA). Samples were spiked with known concentrations (20–100 nM) of mercury(II) ion and analyzed by the developed method. In each sample, the fluorescence quenching of PQ<sup>+</sup> .Cl<sup>−</sup> was immediately measured after spiking of mercury(II) onto the water samples. For method validation, mercury(II) concentrations in the spiked samples were also determined by the standard ICP-OES method. The results are summarized in **Table 5**. The recovery percentage of the measured mercury(II) added to the real samples by the developed and the standard ICP-OES methods was in the range from 97.74 to 104.7% and 98.05 to 102.7%, respectively. The calculated values of the Student's *t* and *F* tests were found lower than the tabulated Student's *t* and *F* tests at 0.05 probability [78] revealing no significant differences between both methods. Thus, it can be concluded that the proposed fluorescence probe can be used as a potential assay for sensing of mercury(II) in complex matrices.

#### **4. Conclusion**

A new and facile extractive spectrofluorometric method for cost effective, precise, accurate and selective determination of trace levels of mercury(II) in water. The proposed method was based upon formation of ternary ion associate complex [(PQ+ ) 2 .(HgI4 )2−] between [HgI4 ]2− and the fluorescence probe PQ<sup>+</sup> .Cl<sup>−</sup> . The proposed system offered excellent selectivity towards mercury(II) ions over most anions and metal cations. The detection process could be performed quickly at room temperature without any catalyst or oxidizer. The proposed method provides LOD (1.3 nM) lower than the value set by WHO (10 nM) and USEPA for drinkable water [1, 17]. The developed method is easy to operate as it does not require sophisticated experimental techniques, and the proposed assay is useful for point-of care applications. Moreover, the method opens capable ways for developing fluorescence assay strategies. The proposed approach was validated successfully by analysis of mercury(II) in environmental water samples by ICP-OES data and statistical treatment of data in terms of significant tests, e.g. *F* and Student's *t* tests. The method could be expanded for ultra-trace analysis of mercury(II) ions in water after on-line enrichment on nanosized solid phase extractor, e.g. polyurethane foam packed column followed by elution with selective reagent [83] and/or its proposed coupling with the advanced microextraction techniques [84]. Therefore, the present work suggested the suitability of the proposed method for use in routine analysis and applicable strategy for analysis of mercury(II) in complex matrices.

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## **Author details**

Dyab A. Al-Eryani1,2, Waqas Ahmad1 , Zeinab M. Saigl<sup>1</sup> , Hassan Alwael1 , Saleh O. Bahaffi<sup>1</sup> , Yousry M. Moustafa<sup>3</sup> and Mohammad S. El-Shahawi1,4†\*

\*Address all correspondence to: mohammad\_el\_shahawi@yahoo.co.uk

1 Department of Chemistry, Faculty of Science, King Abdulaziz University, Jeddah, Saudi Arabia

2 Department of Chemistry, Faculty of Applied Science, Thamar University, Thamar, Yemen

3 Faculty of Applied Science, Umm Al-Qura University, Makkah, Saudi Arabia

4 Department of Chemistry, Faculty of Science, Damiatta University, Damiatta, Egypt

†The author is on sabbatical leave from the Department of Chemistry, Faculty of Science, Damiatta University, Damiatta, Egypt.

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water samples by ICP-OES data and statistical treatment of data in terms of significant tests, e.g. *F* and Student's *t* tests. The method could be expanded for ultra-trace analysis of mercury(II) ions in water after on-line enrichment on nanosized solid phase extractor, e.g. polyurethane foam packed column followed by elution with selective reagent [83] and/or its proposed coupling with the advanced microextraction techniques [84]. Therefore, the present work suggested the suitability of the proposed method for use in routine analysis and appli-

, Zeinab M. Saigl<sup>1</sup>

2 Department of Chemistry, Faculty of Applied Science, Thamar University, Thamar, Yemen

and Mohammad S. El-Shahawi1,4†\*

1 Department of Chemistry, Faculty of Science, King Abdulaziz University, Jeddah,

4 Department of Chemistry, Faculty of Science, Damiatta University, Damiatta, Egypt

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, Hassan Alwael1

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, Saleh O. Bahaffi<sup>1</sup>

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Saudi Arabia

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**Chapter 8**

**Provisional chapter**

**Biomonitoring of Trace Metals in the Coastal Waters**

**Biomonitoring of Trace Metals in the Coastal Waters** 

Several environmental contaminants including toxic trace metals are being discharged into the coastal environment causing serious threat to marine organisms and posing public health risk. Marine bivalves (mussel, oyster, and clam) have been successfully used as sentinel organisms for monitoring contaminant levels, including trace metals, in coastal waters around the globe. Chemical analyses measure the contaminants present in the biota but do not necessarily reveal potential biological effects. Therefore, the need to detect and assess the effects of contaminants, especially at low concentrations, has led to the development of molecular markers of contaminant effects called biomarkers. Owing to their short time of response, biomarkers in marine bivalves are used as early warning signals of biological effects caused by environmental pollutants. Research into the development and application of accurate biomarker-based monitoring tools for the environmental contaminants has been intensified in several developed countries.

**Keywords:** bivalves, bioaccumulation, biomarkers, trace metals, mussel watch

Marine pollution is a major problem that has negative effects on the ocean's ecosystems. Economic developments and urbanization are taking place at an accelerated rate in the coastal zones across the world, putting enormous pressures on coastal waters and marine habitats. Incidents of coastal and marine water pollution have increased throughout the world, mainly due to discharges from rivers, increased surface run-off, drainage from expanding port areas, oil spills, discharges from shipping activities, and domestic and industrial effluent discharges.

> © 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

© 2018 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use,

distribution, and reproduction in any medium, provided the original work is properly cited.

DOI: 10.5772/intechopen.76938

**Using Bivalve Molluscs**

**Using Bivalve Molluscs**

Periyadan K. Krishnakumar,

and Geetha Sasikumar

**Abstract**

**1. Introduction**

http://dx.doi.org/10.5772/intechopen.76938

Mohammad A. Qurban and Geetha Sasikumar

Periyadan K. Krishnakumar, Mohammad A. Qurban

Additional information is available at the end of the chapter

Additional information is available at the end of the chapter


#### **Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs**

DOI: 10.5772/intechopen.76938

Periyadan K. Krishnakumar, Mohammad A. Qurban and Geetha Sasikumar Periyadan K. Krishnakumar, Mohammad A. Qurban and Geetha Sasikumar

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/intechopen.76938

#### **Abstract**

[74] Sun Y, Ma L, Wang HY, Tang B. Determination of procaine hydrochloride by fluorim-

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ed. Belmont, USA: Cengage Learning; 2014

1988;**205**:139-147

152 Trace Elements - Human Health and Environment

2010

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203-211

Acta Part A. 2009;**72**:533-537

Several environmental contaminants including toxic trace metals are being discharged into the coastal environment causing serious threat to marine organisms and posing public health risk. Marine bivalves (mussel, oyster, and clam) have been successfully used as sentinel organisms for monitoring contaminant levels, including trace metals, in coastal waters around the globe. Chemical analyses measure the contaminants present in the biota but do not necessarily reveal potential biological effects. Therefore, the need to detect and assess the effects of contaminants, especially at low concentrations, has led to the development of molecular markers of contaminant effects called biomarkers. Owing to their short time of response, biomarkers in marine bivalves are used as early warning signals of biological effects caused by environmental pollutants. Research into the development and application of accurate biomarker-based monitoring tools for the environmental contaminants has been intensified in several developed countries.

**Keywords:** bivalves, bioaccumulation, biomarkers, trace metals, mussel watch

#### **1. Introduction**

Marine pollution is a major problem that has negative effects on the ocean's ecosystems. Economic developments and urbanization are taking place at an accelerated rate in the coastal zones across the world, putting enormous pressures on coastal waters and marine habitats. Incidents of coastal and marine water pollution have increased throughout the world, mainly due to discharges from rivers, increased surface run-off, drainage from expanding port areas, oil spills, discharges from shipping activities, and domestic and industrial effluent discharges.

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2018 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

Most of the world's wastes around 20 billion tons per year end up in the sea, often without any preliminary processing.

Trace metals are introduced into the coastal waters through natural process and anthropogenic activities. The natural process includes river discharge, rock weathering, wind-generated dust from arid and semi-arid regions of the continents, and hydrothermal circulation at mid-ocean ridges. The anthropogenic sources of metals include agriculture, fossil fuel extraction, refining and burning, chemical production, and intentional and accidental discharges. Trace levels of trace metals naturally occur in the marine environment, and many of them at low concentrations are essential for marine life. However, if their concentrations exceed the natural levels, it will cause a serious threat to marine life. Monitoring and assessment programs are routinely conducted in the coastal waters for planning and implementing mitigation measures to control trace metal pollution. Historically as one of the simple and widely used monitoring techniques, sampling, and analysis of seawater and sediment are being employed for estimating the levels of contaminants including trace metals in coastal waters. Instead of using water or sediment samples, tissue concentrations of contaminants in marine organisms, especially bivalves, are being used as a reliable method for assessing the coastal water quality since 1960s [1–4].

Most of the marine bivalves such as mussels, oysters, and clams are commercially important groups, and several of them are being used for coastal farming around the globe and as popular seafood. Since late 1960s and early 1970s, bivalves such as mussels were used for biomonitoring trace metals in coastal waters [3, 5]. In biomonitoring, tissue burden of trace metals in marine organisms are analyzed, and the biological responses of organisms are measured to assess changes in the environmental quality caused by toxic contaminants [6–8]. This chapter will attempt to provide an overview of the basic concept, methods and the present status of the biomonitoring of trace metals in the coastal waters using bivalve molluscs.

> range of contaminants and may thrive even in highly polluted environments [3, 14]. These qualities make them a group of candidate species for biomonitoring programs across the globe. As filter feeders, they bioaccumulate various contaminants and their tissue concentrations provide a time-integrated picture of contaminants in the environment [15, 16]. It has been reported that bivalves accumulate trace metals in their tissues at levels up to 100–100,000 times higher than the concentrations observed in the seawater in which they live [5, 17]. Therefore, several chemical contaminants, including trace metals, present at undetectable levels in seawater can be detected in bivalve tissues. Different species of clams, mussels, and oysters have widespread distribution across the continents (**Figure 1**), and many of those species have been successfully

> **Figure 1.** Common marine bivalves and their habitats from the Indian coast. (A) Intertidal rocky area showing green mussel beds from the south west coast of India; (B) green mussel *Perna viridis*; (C) enlarged view of green mussels; (D) oyster bed consisting of *Crassostrea madrasensis* and *Saccostrea cucullata* exposed during low tide; and (E) enlarged view of *C*. *madrasensis and S. cucullata;* (E) clam bed consisting of *Meretrix casta* and (F) enlarged view of the clam *Paphia malabarica*.

Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs

http://dx.doi.org/10.5772/intechopen.76938

155

used for monitoring the concentrations of contaminants in the marine environment [5].

Cobalt, copper, chromium, iron, magnesium, manganese, molybdenum, nickel, selenium, and zinc are essential metals that are required for various biochemical and physiological functions of animals [18] while other metals such as aluminum, antinomy, arsenic, barium, cadmium, gold, lead, lithium, mercury, nickel, platinum, silver, strontium, tin, titanium, and vanadium have no

**3. Metal bioaccumulation in bivalves**

## **2. Why bivalves**

Generally, bivalves are suspension feeders or deposit feeders, or even utilize both feeding methods. They feed on microscopic algae, bacteria, and detritus through filter feeding process. They draw water from the posterior ventral side through the inhalant siphon, and the water passes through the gills and gets expelled through the exhalent siphon. In this process, they filter large quantities of seawater, and the water filtering capacity of typical natural mussel beds has been calculated as 7–12 m<sup>3</sup> , m−1, h−1 [9, 10]. One single adult blue mussel pumps around 50 ml of seawater per minute during active feeding [11]. As bivalves filter large quantiles of seawater, their tissues absorb some of the contaminants present in water and food particles. Bivalves accumulate trace metals from the surrounding aquatic medium across the cellular membrane (dissolved source) and from the food materials (dietary source) [12].

Historically, bivalve molluscs are considered as valuable marine organisms for environmental monitoring and used as biomonitors of chemical pollution of coastal waters [3, 5, 13]. Bivalves are widely distributed from the North Pole to the South Pole, sessile in nature, and easy to sample and available in a suitable size for chemical analysis. Bivalves are also resistant to a wide

**Figure 1.** Common marine bivalves and their habitats from the Indian coast. (A) Intertidal rocky area showing green mussel beds from the south west coast of India; (B) green mussel *Perna viridis*; (C) enlarged view of green mussels; (D) oyster bed consisting of *Crassostrea madrasensis* and *Saccostrea cucullata* exposed during low tide; and (E) enlarged view of *C*. *madrasensis and S. cucullata;* (E) clam bed consisting of *Meretrix casta* and (F) enlarged view of the clam *Paphia malabarica*.

range of contaminants and may thrive even in highly polluted environments [3, 14]. These qualities make them a group of candidate species for biomonitoring programs across the globe. As filter feeders, they bioaccumulate various contaminants and their tissue concentrations provide a time-integrated picture of contaminants in the environment [15, 16]. It has been reported that bivalves accumulate trace metals in their tissues at levels up to 100–100,000 times higher than the concentrations observed in the seawater in which they live [5, 17]. Therefore, several chemical contaminants, including trace metals, present at undetectable levels in seawater can be detected in bivalve tissues. Different species of clams, mussels, and oysters have widespread distribution across the continents (**Figure 1**), and many of those species have been successfully used for monitoring the concentrations of contaminants in the marine environment [5].

## **3. Metal bioaccumulation in bivalves**

Most of the world's wastes around 20 billion tons per year end up in the sea, often without

Trace metals are introduced into the coastal waters through natural process and anthropogenic activities. The natural process includes river discharge, rock weathering, wind-generated dust from arid and semi-arid regions of the continents, and hydrothermal circulation at mid-ocean ridges. The anthropogenic sources of metals include agriculture, fossil fuel extraction, refining and burning, chemical production, and intentional and accidental discharges. Trace levels of trace metals naturally occur in the marine environment, and many of them at low concentrations are essential for marine life. However, if their concentrations exceed the natural levels, it will cause a serious threat to marine life. Monitoring and assessment programs are routinely conducted in the coastal waters for planning and implementing mitigation measures to control trace metal pollution. Historically as one of the simple and widely used monitoring techniques, sampling, and analysis of seawater and sediment are being employed for estimating the levels of contaminants including trace metals in coastal waters. Instead of using water or sediment samples, tissue concentrations of contaminants in marine organisms, especially bivalves, are being used as a reliable method for assessing the coastal water quality since 1960s [1–4].

Most of the marine bivalves such as mussels, oysters, and clams are commercially important groups, and several of them are being used for coastal farming around the globe and as popular seafood. Since late 1960s and early 1970s, bivalves such as mussels were used for biomonitoring trace metals in coastal waters [3, 5]. In biomonitoring, tissue burden of trace metals in marine organisms are analyzed, and the biological responses of organisms are measured to assess changes in the environmental quality caused by toxic contaminants [6–8]. This chapter will attempt to provide an overview of the basic concept, methods and the present status of

Generally, bivalves are suspension feeders or deposit feeders, or even utilize both feeding methods. They feed on microscopic algae, bacteria, and detritus through filter feeding process. They draw water from the posterior ventral side through the inhalant siphon, and the water passes through the gills and gets expelled through the exhalent siphon. In this process, they filter large quantities of seawater, and the water filtering capacity of typical natural mus-

around 50 ml of seawater per minute during active feeding [11]. As bivalves filter large quantiles of seawater, their tissues absorb some of the contaminants present in water and food particles. Bivalves accumulate trace metals from the surrounding aquatic medium across the cellular membrane (dissolved source) and from the food materials (dietary source) [12].

Historically, bivalve molluscs are considered as valuable marine organisms for environmental monitoring and used as biomonitors of chemical pollution of coastal waters [3, 5, 13]. Bivalves are widely distributed from the North Pole to the South Pole, sessile in nature, and easy to sample and available in a suitable size for chemical analysis. Bivalves are also resistant to a wide

, m−1, h−1 [9, 10]. One single adult blue mussel pumps

the biomonitoring of trace metals in the coastal waters using bivalve molluscs.

any preliminary processing.

154 Trace Elements - Human Health and Environment

**2. Why bivalves**

sel beds has been calculated as 7–12 m<sup>3</sup>

Cobalt, copper, chromium, iron, magnesium, manganese, molybdenum, nickel, selenium, and zinc are essential metals that are required for various biochemical and physiological functions of animals [18] while other metals such as aluminum, antinomy, arsenic, barium, cadmium, gold, lead, lithium, mercury, nickel, platinum, silver, strontium, tin, titanium, and vanadium have no established biological functions and are considered as non-essential metals [19]. However, the essential metals will be harmful to the organisms if their concentrations exceed the natural levels. The expert's group of International Council for the Exploration of the Sea (ICES) and Oslo and Paris Conventions (OSPAR) highlighted the trace metals such as arsenic, cadmium, chromium, copper, mercury, nickel, lead, and zinc in the marine environment as key substances of concern [20].

Bivalves accumulate both essential and non-essential metals in their soft tissues above the background levels in seawater or sediments, and this process is called bioaccumulation. Bioaccumulation is a good integrative indicator of the chemical exposures of marine organisms such as bivalves in polluted waters [21]. Trace metals cannot be metabolized by organisms, and hence bioaccumulation of trace metals is of particular value as an exposure indicator. However, metal bioaccumulation can be complex. The bioaccumulation levels in mollusks differ among metals in the same bivalve species and among species [13, 21–23] due to the biological role of different metals and to specific strategies of accumulation [23]. In addition, the metal bioaccumulation in bivalves depends on the marine environmental factors (temperature, pH, salinity, co-occurrence of metals, etc.) and the biological conditions (age, sex, sexual maturity stage, etc.) of the species [24, 25].

The gill tissue of bivalves constitutes a key interface for the uptake of dissolved metal ions from water followed by the mantle tissue, and the uptake of metals bound to particulate material is achieved via the digestive tract, in particular, via the digestive gland [23]. Generally, in bivalves, maximum concentrations of metals have been reported in the digestive gland and/or gill tissue followed by mantle and muscle tissue [26, 27]. The bioaccumulation of trace metals in bivalve tissues is dependent on different metabolic processes occurring within specific cell types in target tissues. Metallothioneins (MTs), the low-molecular-weight proteins present in organisms including bivalves are involved in the intracellular regulation of metals such as Cu, Zn, and Cd [28]. Epithelial cells of gill and mantle can synthesize MT and sequester metals into the lysosomes for further transport in circulating hemocytes [29].

## **4. Bivalves as sentinel organisms**

*Sentinel organisms* accumulate contaminants in their tissues without any harmful effects and can be measured in a sensitive manner the amount of contaminants that are biologically available [30]. Several comprehensive reviews have been published on the use of bivalve molluscs as sentinel organisms and as biomonitors of metal pollution [5, 12, 20, 31–35]. These reviews and studies provide an in-depth discussion on metal bioaccumulation and metal bioavailability, highlighting the historical usage of bivalves in environmental studies.

Most of the bivalves such as clams, mussels, and oysters, fulfill the criteria required for a typical sentinel organisms and being successfully used as spatial and temporal trend indicators of contaminants in monitoring program from several parts of the world [3, 7, 12, 14–16, 36–39]. The tissue concentrations of various toxic trace metals in wild mussel species from various regions worldwide are summarized in **Table 1**. The tissue concentrations ranged from low to high values depending upon the environmental status of the study area.

**Country**

San

*Mytilus edulis*

6.9

4.05

mg/kg dry wt

Francisco

Bay, USA

Claisebrook

*Xenostrobus* 

12–61 0.46–

0.21–

0.05–

0.06–

1.7–

<0.01

0.22–

0.08–

6–9.6

0.34–

3.3–

[99]

28

0.57

85

0.52

0.75

0.27

0.17

0.16

2.2

*sp.* mg/kg wet

Cove,

Western

wt

Australia

South

*Perna* 

5.35–

0.14–

2.19–

0.08–

0.13–

45.31–

[100]

5.83

1.53

147.18

18.25

27.48

1.67

Island New

*canaliculatus*

mg/kg dry

wt

Offshore

*Bathymodiolus* 

2.6–

2.16–

5.89–

0.78–

0.81–

0.1–

5.53–

14.28–

0.42–

4.41–

33.76–

17.98–

2.69–

4.5–

[101]

45.78

4.06

8.01

[102]

1.25

4.8

79.04

*platifrons* mg/

25.13

6.73

10.03

4.35

1.72

0.45

42.31

56.07

kg dry wt

South China

Sea

East coast of

*Perna viridis*

0.01–

12.64–

0.48–

1.51–

1.45–

96.62–

1.3–

0.44–

66.05–

4.78

2.93

231

20.95

4–30

1–2.9

3.7–

53.4–

0.8–5 2–7

59.1–

2–13

[103]

273

11.1

4.6–

128–

132–

7.3–

[104]

Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs

85.0

345

17.2

5.5–

123–

[105]

http://dx.doi.org/10.5772/intechopen.76938

180

11.5

6.7–

120–

208–

4.5–

[106]

11.7

157

320

9.5

415

603

719

5.31

10.93

28.55

1002

mg/kg dry wt

0.14

China

East Adriatic

*Mytilus* 

Sea, Croatia

*galloprovincialis* mg/kg

dry wt

Adriatic Sea

*Mytilus* 

(Montenegro

*galloprovincialis* mg/kg dry wt

Tyrrhenian

*Mytilus* 

Sea (Gulf of

*galloprovincialis* mg/kg

Gaeta)

dry wt

Marmara

*Mytilus* 

*galloprovincialis* mg/kg

Sea (NW

coasts)

dry wt

coasts)

Zealand

**Mussel** 

**Ag**

**Al**

**As**

**Cd**

**Cr**

**Co**

**Cu**

**Fe**

**Hg**

**Ni**

**Pb**

**Zn**

92

4.6

**Ti Se**

**V**

**Sr**

**Ba**

**Mn**

**Ref.**

[97,

98]

**Species**


established biological functions and are considered as non-essential metals [19]. However, the essential metals will be harmful to the organisms if their concentrations exceed the natural levels. The expert's group of International Council for the Exploration of the Sea (ICES) and Oslo and Paris Conventions (OSPAR) highlighted the trace metals such as arsenic, cadmium, chromium, copper, mercury, nickel, lead, and zinc in the marine environment as key substances of concern [20].

Bivalves accumulate both essential and non-essential metals in their soft tissues above the background levels in seawater or sediments, and this process is called bioaccumulation. Bioaccumulation is a good integrative indicator of the chemical exposures of marine organisms such as bivalves in polluted waters [21]. Trace metals cannot be metabolized by organisms, and hence bioaccumulation of trace metals is of particular value as an exposure indicator. However, metal bioaccumulation can be complex. The bioaccumulation levels in mollusks differ among metals in the same bivalve species and among species [13, 21–23] due to the biological role of different metals and to specific strategies of accumulation [23]. In addition, the metal bioaccumulation in bivalves depends on the marine environmental factors (temperature, pH, salinity, co-occurrence of metals, etc.) and the biological conditions (age,

The gill tissue of bivalves constitutes a key interface for the uptake of dissolved metal ions from water followed by the mantle tissue, and the uptake of metals bound to particulate material is achieved via the digestive tract, in particular, via the digestive gland [23]. Generally, in bivalves, maximum concentrations of metals have been reported in the digestive gland and/or gill tissue followed by mantle and muscle tissue [26, 27]. The bioaccumulation of trace metals in bivalve tissues is dependent on different metabolic processes occurring within specific cell types in target tissues. Metallothioneins (MTs), the low-molecular-weight proteins present in organisms including bivalves are involved in the intracellular regulation of metals such as Cu, Zn, and Cd [28]. Epithelial cells of gill and mantle can synthesize MT and sequester metals

*Sentinel organisms* accumulate contaminants in their tissues without any harmful effects and can be measured in a sensitive manner the amount of contaminants that are biologically available [30]. Several comprehensive reviews have been published on the use of bivalve molluscs as sentinel organisms and as biomonitors of metal pollution [5, 12, 20, 31–35]. These reviews and studies provide an in-depth discussion on metal bioaccumulation and metal bioavail-

Most of the bivalves such as clams, mussels, and oysters, fulfill the criteria required for a typical sentinel organisms and being successfully used as spatial and temporal trend indicators of contaminants in monitoring program from several parts of the world [3, 7, 12, 14–16, 36–39]. The tissue concentrations of various toxic trace metals in wild mussel species from various regions worldwide are summarized in **Table 1**. The tissue concentrations ranged from low to

sex, sexual maturity stage, etc.) of the species [24, 25].

156 Trace Elements - Human Health and Environment

**4. Bivalves as sentinel organisms**

into the lysosomes for further transport in circulating hemocytes [29].

ability, highlighting the historical usage of bivalves in environmental studies.

high values depending upon the environmental status of the study area.

Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs http://dx.doi.org/10.5772/intechopen.76938 157


**Country**

Turkey

*Mytilus* 

*galloprovincialis*

mg/kg *dry* 

*weight*

Eastern

Aegean Sea

Italy Venice

*Mytilus* 

1.16–

0.16–

3.55–

1.08–

135–

[115]

4.27

400

6.59

2.75

10.8

*galloprovincialis*

mg/kg *dry* 

*weight*

Brazil

*Mytella* 

778–

1.44–

Bdl–

Bdl–

Bdl–

6.03–

Bdl–

Bdl–

Bdl–

50.8–

Bdl–

Bdl–

35.5–

Bdl–

30.7–

[116]

2458

23.1

1.42

3.13

611

102

1820

0.35

19.4

141

49.6

6.93

95.8

88.7

3520

*guyanensis*

mg/kg *dry* 

*weight*

India

*Perna viridis*

0.24–

Bdl–

Bdl–

Bdl–

Bdl–

Bdl–

Bdl–

1.91–

[15]

8.77

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159

2.89

1.95

17.36

3.49

0.46

Selected trace metal concentrations in the soft tissue of wild mussel species from various regions worldwide.

1.84

235.6

mg/kg *wet* 

*weight*

**Table 1.**

Lagoon

**Mussel** 

**Ag**

**Al**

**As**

**Cd** 0.24–

0.32–

2.44–

0.11–

0.84–

75.9–

0.15

2.41

201

0.49

7.27

5.49

**Cr**

**Co**

**Cu**

**Fe**

**Hg**

**Ni**

**Pb**

**Zn**

**Ti Se**

**V**

**Sr**

**Ba**

**Mn**

**Ref.**

[114]

**Species**


**Country**

**Mussel** 

**Ag**

**Al**

**As**

**Cd**

**Cr**

**Co**

**Cu** 3.5–

48.6–

17.8–

28.5

5.3

3.9–

159–

351

9.7

1.3–

11.0–

13.3–

[109]

15.2

1.8

11.7

49.9

**Fe**

**Hg**

**Ni**

**Pb**

**Zn**

**Ti Se**

**V**

**Sr**

**Ba**

**Mn** 2.6–

[107]

4.7

[108]

158 Trace Elements - Human Health and Environment

**Ref.**

**Species**

Aegean Sea

*Mytilus* 

*galloprovincialis* 

*mg/kg dry wt*

N Atlantic

*Mytilus* 

(Spanish

*galloprovincialis* 

*mg/kg dry wt*

Gallician

coasts)

Island of

*Mytilus* 

Gossa (W

*galloprovincialis* 

*mg/kg dry wt*

coast of

Norway)

Spain

*Mytilus* 

14.6–

0.4–

2.6–

0.4–

9.1–

1.5–

1.1–

202.7–

5.8–

1.7–

5.6–

[110]

55.3

15.4

13.3

300.8

8.7

7.1

31.5

2.3

5.7

69.3

34.8

0.7–

24.3–

43.8–

0.4–

[111]

4.8

[112]

133.5

12.9

82.0

*galloprovincialis*

mg/kg *dry* 

*weight*

N Aegean

*Mytilus* 

Sea (Strait of

*galloprovincialis* 

Canakkale)

Trinidad

*Perna viridis*

0.01–

0.06–

1.02–

0.03–

0.3–

11.3–

40.37

0.07

0.75

0.61

0.2

1.98

mg/kg *wet* 

*weight*

Venezuela

*Perna viridis*

0.02–

0.12–

1.42–

0.02–

0.22–

8.75–

[112]

16.38

0.08

1.3

0.05

0.33–

0.46–

5.51–

1.67–

123–

[105]

2.49

180

0.49

1.31

11.5

11.7–

312–

46.9–

[113]

73.0

396

23.3

0.16

3.43

mg/kg *wet* 

*weight*

Italy

*Mytilus* 

Tyrrhenian

*galloprovincialis*

coastal areas

mg/kg *dry* 

*weight*

Black Sea

*Mytilus* 

*galloprovincialis* 

*mg/kg dry wt*

(Turkish

coasts)

*mg/kg dry wt*

Cantabrian

Coast

**Table 1.** Selected trace metal concentrations in the soft tissue of wild mussel species from various regions worldwide.

#### Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs http://dx.doi.org/10.5772/intechopen.76938 159

#### **4.1. Mussel watch programs**

Mussels and other marine bivalves are widely used as sentinel organisms in "mussel watch" programs for indicating levels of pollutants in the coastal marine environment due to their ability to bioaccumulate organic or toxic elements [40]. Under mussel watch program, environmental contaminants (trace metals, hydrocarbons, pesticides, etc.) accumulated in the soft tissue of natural, cultured, or deployed bivalves (clams, mussels, and oysters) collected from a set of defined geographical locations over a time-span of several years are systematically and repeatedly measured for assessing and comparing the coastal water quality [5, 40–42]. A prominent example is the US Mussel Watch Program originally started in 1976 [3, 43] and established as the Mussel Watch component of National Oceanic and Atmospheric Administration's (NOAA) National Status and Trends (NST) program during 1986–2012 [44, 45, 46]. In spite of the criticisms and limitations [47], the US mussel watch results made valuable contributions to our understanding of trace metal contamination and its biogeochemistry in coastal ecosystems [5].

Later, the contaminant monitoring programs similar to mussel watch were implemented throughout the world either for monitoring long-term spatial and temporal pollution trends covering large marine region containing multiple monitoring stations and several anthropogenic contamination sources [36–38, 48–51] or for monitoring and solving local pollution

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161

The mussel watch program initiated in USA has led to the formation of the International Mussel Watch (IMW) Projects [5]. It was initiated by the International Oceanographic Commission (IOC) in collaboration with the United Nations Environment Program (UNEP) and the US NOAA. **Table 2** summarizes the details of the international mussel watch program conducted from different geographical locations. Recently, the advantages and limitations of

Chemical analyses of bivalve tissue samples measure the contaminants present but do not necessarily reveal potential biological effects on bivalves. Therefore, biomarkers were developed to assess the health status of the marine organisms, especially bivalves. Biomarkers are the early warning signals about the health status of bivalves exposed to toxic contaminants, because a toxic effect or response will be apparent at the molecular or cellular level before it is noticeable at higher biological levels. The concept of biomarker is borrowed from medical science, which describes a measurable indicator such as blood cholesterol profile connected to relevant clinical endpoints like atherosclerosis and heart attack. The biochemical biomarkers (acetylcholinesterase inhibition for exposure to neurotoxic compounds, cytochrome P450 for detoxification of polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs), and the different methods to detect genotoxicity), which are used in marine environ-

During the last decade, several biomarkers sensitive to contaminant exposure and/or impact have been developed as tools for use in marine environmental monitoring and impact assessment [7, 8, 62]. During the same time, various monitoring agencies began to focus on locating the source of contamination and fates as well as the impact as contaminants are usually discharged into the coastal waters, especially estuaries, where effects have been most significantly detected. The European Union's Water Framework Directive (WFD) also stressed the requirement of monitoring programs to assess the achievement of good chemical and ecological status for all water bodies by 2015 [63]. In the past 30–40 years, numerous biomarkers have been developed on bivalve mollusks, especially mussels (see **Table 3**) with the objective to apply them for environmental biomonitoring. Biomarkers based on responses at physiological level, cellular/tissue level, and molecular level of bivalve molluscs are developed and recommended as tools for studying the effects of contaminants on field and laboratory exposed bivalves, especially mussels [6, 64–66]. Research into the development and application of accurate biomarker-based monitoring tools for the environmental contaminants has been intensified in several developed countries, and they are using several biomarkers based in marine bivalves to monitor the environmental quality of coastal and estuarine waters [20].

problems covering a small geographical areas [7, 8, 15, 32, 52–58].

the mussel watch concept were discussed 40 years after its inception [5].

**5. Biomarkers of exposure in bivalves**

mental monitoring are still used in humans [59–61].


**Table 2.** Details of the International Mussel Watch (IMW) program conducted from various parts of the globe [5].

Later, the contaminant monitoring programs similar to mussel watch were implemented throughout the world either for monitoring long-term spatial and temporal pollution trends covering large marine region containing multiple monitoring stations and several anthropogenic contamination sources [36–38, 48–51] or for monitoring and solving local pollution problems covering a small geographical areas [7, 8, 15, 32, 52–58].

The mussel watch program initiated in USA has led to the formation of the International Mussel Watch (IMW) Projects [5]. It was initiated by the International Oceanographic Commission (IOC) in collaboration with the United Nations Environment Program (UNEP) and the US NOAA. **Table 2** summarizes the details of the international mussel watch program conducted from different geographical locations. Recently, the advantages and limitations of the mussel watch concept were discussed 40 years after its inception [5].

## **5. Biomarkers of exposure in bivalves**

**4.1. Mussel watch programs**

160 Trace Elements - Human Health and Environment

IMW Phase I (Initial Implementation): 1991–1993

IMW Phase II 1997–1999

IMW Pilot Study— Black Sea. 1996–1997

Mediterranean Basin and the International Mediterranean Commission (CISEM) Mussel Watch program. 2002–2006

Western

South America, Central America, Mexico and Caribbean

Asia Pacific Region (Japan, South Korea, Russia, China, the Philippines, Vietnam, Malaysia, Cambodia, Thailand, Indonesia and India)

Six Black Sea Countries (Bulgaria, Georgia, Romania, Russia, Turkey and Ukraine).

The coasts of the Western Mediterranean Basin (Spain, France, Italy, North Tunisia, Algeria and Morocco)

Mussels and other marine bivalves are widely used as sentinel organisms in "mussel watch" programs for indicating levels of pollutants in the coastal marine environment due to their ability to bioaccumulate organic or toxic elements [40]. Under mussel watch program, environmental contaminants (trace metals, hydrocarbons, pesticides, etc.) accumulated in the soft tissue of natural, cultured, or deployed bivalves (clams, mussels, and oysters) collected from a set of defined geographical locations over a time-span of several years are systematically and repeatedly measured for assessing and comparing the coastal water quality [5, 40–42]. A prominent example is the US Mussel Watch Program originally started in 1976 [3, 43] and established as the Mussel Watch component of National Oceanic and Atmospheric Administration's (NOAA) National Status and Trends (NST) program during 1986–2012 [44, 45, 46]. In spite of the criticisms and limitations [47], the US mussel watch results made valuable contributions to our understanding of trace metal contamination and its biogeochemistry in coastal ecosystems [5].

**Project phase and year Study areas Bivalve species List of contaminants References**

Total Polychlorinated biphenyls (PCBs), total Chlordane (CHLs),

Total PCBs, dichloro diphenyl trichloroethane and its metabolites (DDTs), CHLs, hexachlorocyclohexane isomers (HCHs) and hexachlorobenzene (HCB), polychlorinated dibenzo-p-dioxins and furans (PCDDs/Fs), coplanar PCBs (Co-PCBs), Butyltins (BTs) and

PAHs, PCBs, DDTs [122]

[5, 117]

[38, 118–121]

[123–125]

and total HCHs

some heavy metals

Heavy metals, chlorinated pesticides and PCBs and PAHs

Blue mussels (*Mytilus* sp.) 134 stations Oysters (*Crassostrea* sp.)–18 stations Other bivalves–24

stations

*viridis*).

sites

sites

Blue mussel, (*M. edulis*), and the green mussel (*Perna* 

Blue mussels (*M. galloprovincialis*)- 5–13

Caged mussels (*Mytilus* sp.) deployed at 122

**Table 2.** Details of the International Mussel Watch (IMW) program conducted from various parts of the globe [5].

Chemical analyses of bivalve tissue samples measure the contaminants present but do not necessarily reveal potential biological effects on bivalves. Therefore, biomarkers were developed to assess the health status of the marine organisms, especially bivalves. Biomarkers are the early warning signals about the health status of bivalves exposed to toxic contaminants, because a toxic effect or response will be apparent at the molecular or cellular level before it is noticeable at higher biological levels. The concept of biomarker is borrowed from medical science, which describes a measurable indicator such as blood cholesterol profile connected to relevant clinical endpoints like atherosclerosis and heart attack. The biochemical biomarkers (acetylcholinesterase inhibition for exposure to neurotoxic compounds, cytochrome P450 for detoxification of polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs), and the different methods to detect genotoxicity), which are used in marine environmental monitoring are still used in humans [59–61].

During the last decade, several biomarkers sensitive to contaminant exposure and/or impact have been developed as tools for use in marine environmental monitoring and impact assessment [7, 8, 62]. During the same time, various monitoring agencies began to focus on locating the source of contamination and fates as well as the impact as contaminants are usually discharged into the coastal waters, especially estuaries, where effects have been most significantly detected. The European Union's Water Framework Directive (WFD) also stressed the requirement of monitoring programs to assess the achievement of good chemical and ecological status for all water bodies by 2015 [63]. In the past 30–40 years, numerous biomarkers have been developed on bivalve mollusks, especially mussels (see **Table 3**) with the objective to apply them for environmental biomonitoring. Biomarkers based on responses at physiological level, cellular/tissue level, and molecular level of bivalve molluscs are developed and recommended as tools for studying the effects of contaminants on field and laboratory exposed bivalves, especially mussels [6, 64–66]. Research into the development and application of accurate biomarker-based monitoring tools for the environmental contaminants has been intensified in several developed countries, and they are using several biomarkers based in marine bivalves to monitor the environmental quality of coastal and estuarine waters [20].


**5.2. Cellular biomarkers**

the LMS in mussels as one of the core biomarkers [86–88].

monitoring studies [67, 68, 85, 90].

**5.3. Biomarkers of genotoxicity**

breaks in marine bivalves.

The digestive gland cells in bivalves play a key role in digestive and absorptive processes and also in the detoxification and excretion of contaminants [80]. The lysosomal system in the digestive cells was identified as the main target site for the toxic effects of most of the environmental contaminants including trace metals [81]. Lysosomal responses to cell injury due to contaminant exposure or stress caused by environmental changes fall into three categories: (1) changes in lysosomal contents, (2) changes in fusion events, and (3) changes in membrane permeability [81]. Changes in lysosomal membrane permeability of bivalves can be measured using the lysosomal membrane stability (LMS) test [82–84]. The LMS test can be conducted by using two different methodologies: (i) a cytochemical method using cryostat sections of digestive gland tissue and (ii) an in vivo cytochemical method using hemolymph cells. Biomarkers such as LMS, accumulation of lipofuscin and neutral lipids in bivalves were successfully used for coastal pollution monitoring studies [7, 8, 69, 70, 82–84]. Subsequently, different regional conventions have recommended the use of LMS as a general stress biomarker of chemical pollution within the framework of the pollution biomonitoring programs [67, 68, 85]. The proposed integrated assessment approach of contaminants and their effects in the NE Atlantic Baltic Sea Action Plan and in the Mediterranean Ecosystem Approach (EcAp) have included

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It has been demonstrated that metallothioneins (inducible low molecular, sulfhydryl proteins) levels in the digestive cells of bivalves will be induced after exposure to trace metals such as Cd, Cu, and Zn [89]. The induction of metallothioneins (MT) in bivalves has been proposed as biomarkers of trace metal stress, and it has been recommended to use in coastal pollution

A wide variety of chemical contaminants capable of directly or indirectly damaging the DNA of marine organisms are being discharged into the marine environment. These genotoxic chemicals are capable of inducing some changes in the molecular and cellular levels of marine bivalves [91, 92]. Two well-known tests, micronucleus assay and comet assay, are being widely used to assess the genotoxic effects of environmental contaminants on marine bivalves [91, 92]. The micronucleus assay is used to detect the structural and numerical chromosomal changes while the comet assay (single-cell gel electrophoresis) is used to detect DNA strand

The biomarkers in marine bivalves based on sub-lethal effects of contaminants are ecologically relevant and can be used to give subtle signals of response to contaminants before damage becomes irreversible. The water quality in European coastal sites was classified ranging from class 1 (clean areas) to class 5 (highly polluted areas), based on global biomarker index for

**6. Coastal pollution monitoring using biomarkers a case study**

**Table 3.** List of biomarkers routinely used for monitoring the coastal waters quality using marine bivalves.

#### **5.1. Physiological biomarkers**

The biological indicators of health in bivalves such as Body Condition Index (BCI), stress on stress response (SOS), and scope for growth (SCF) have been recommended as broad markers of stress caused by either environmental changes or contaminants [59, 64, 67–70]. The stress on stress response is a simple test, which measures the mortality rate (time to kill 50% of the sample) of bivalves when exposed to air [70, 71]. The SOS test examines whether stress caused by environmental changes or contaminants have altered the capacity of bivalves to survive under adverse conditions such as aerial exposure. The body condition index (ratio between soft tissue dry/wet weights to its overall size) is a general indicator of favorable growth conditions as well as the overall biological status. The body condition index is routinely used in aquaculture and environmental monitoring studies to assess the health condition of mussels [7, 25, 72].

The growth, reproduction, and survival of bivalves depend on the availability of sufficient energy reserve in their body. Exposure to contaminants negatively affects the energy balance of bivalves due to the high-energy demand for maintaining homeostasis at the expense of growth, storage, defense, and reproduction [73]. Fitness of an individual organism can be measured in terms of Scope for Growth (SfG), which is the measurement of physiological energy balance and it ranges from optimal (positive values) to stressed conditions (negative values) when the organism is exposed to contaminants or unfavorable environmental conditions [74, 75]. The SFG has been widely used in field monitoring studies [76, 77]. The SFG and the growth rates of mussels were drastically reduced when mussels from uncontaminated sites were transplanted along known pollution gradients or placed in the most contaminated areas [78, 79].

#### **5.2. Cellular biomarkers**

The digestive gland cells in bivalves play a key role in digestive and absorptive processes and also in the detoxification and excretion of contaminants [80]. The lysosomal system in the digestive cells was identified as the main target site for the toxic effects of most of the environmental contaminants including trace metals [81]. Lysosomal responses to cell injury due to contaminant exposure or stress caused by environmental changes fall into three categories: (1) changes in lysosomal contents, (2) changes in fusion events, and (3) changes in membrane permeability [81].

Changes in lysosomal membrane permeability of bivalves can be measured using the lysosomal membrane stability (LMS) test [82–84]. The LMS test can be conducted by using two different methodologies: (i) a cytochemical method using cryostat sections of digestive gland tissue and (ii) an in vivo cytochemical method using hemolymph cells. Biomarkers such as LMS, accumulation of lipofuscin and neutral lipids in bivalves were successfully used for coastal pollution monitoring studies [7, 8, 69, 70, 82–84]. Subsequently, different regional conventions have recommended the use of LMS as a general stress biomarker of chemical pollution within the framework of the pollution biomonitoring programs [67, 68, 85]. The proposed integrated assessment approach of contaminants and their effects in the NE Atlantic Baltic Sea Action Plan and in the Mediterranean Ecosystem Approach (EcAp) have included the LMS in mussels as one of the core biomarkers [86–88].

It has been demonstrated that metallothioneins (inducible low molecular, sulfhydryl proteins) levels in the digestive cells of bivalves will be induced after exposure to trace metals such as Cd, Cu, and Zn [89]. The induction of metallothioneins (MT) in bivalves has been proposed as biomarkers of trace metal stress, and it has been recommended to use in coastal pollution monitoring studies [67, 68, 85, 90].

#### **5.3. Biomarkers of genotoxicity**

**5.1. Physiological biomarkers**

Cellular Responses Lysosomal membrane stability

162 Trace Elements - Human Health and Environment

(LMS); lipofuscin and neutral lipids accumulation

Metal-binding cysteine-rich proteins

DNA integrity markers

The biological indicators of health in bivalves such as Body Condition Index (BCI), stress on stress response (SOS), and scope for growth (SCF) have been recommended as broad markers of stress caused by either environmental changes or contaminants [59, 64, 67–70]. The stress on stress response is a simple test, which measures the mortality rate (time to kill 50% of the sample) of bivalves when exposed to air [70, 71]. The SOS test examines whether stress caused by environmental changes or contaminants have altered the capacity of bivalves to survive under adverse conditions such as aerial exposure. The body condition index (ratio between soft tissue dry/wet weights to its overall size) is a general indicator of favorable growth conditions as well as the overall biological status. The body condition index is routinely used in aquaculture and

**Group Biomarker name Description References**

shell length

Stress on stress response (SOS) Assessment of survival rate during aerial exposure

Scope for growth (SFG) Measurement of physiological energy balance

Metallothioneins (MTs) Measurement of metal binding proteins

Micronuclei Assessment of toxic impact on

**Table 3.** List of biomarkers routinely used for monitoring the coastal waters quality using marine bivalves.

comparison with shell cavity volume or

in tissue samples. Compensatory mechanism during exposure to heavy

Assessment of the condition of lysosomes

metals (Cd, Fe, Hg, Zn, As)

and the related cell injury

chromosomes

DNA adducts DNA damage assessment [91, 92, 128] Comet assay Single cell DNA damage assessment [91, 92, 128]

[7, 59, 126]

[71]

[28]

[59, 76]

[7, 8, 61]

[91, 92, 127]

Bivalve Physiology Body Condition Index (BCI) Assessment of tissue weight in

environmental monitoring studies to assess the health condition of mussels [7, 25, 72].

along known pollution gradients or placed in the most contaminated areas [78, 79].

The growth, reproduction, and survival of bivalves depend on the availability of sufficient energy reserve in their body. Exposure to contaminants negatively affects the energy balance of bivalves due to the high-energy demand for maintaining homeostasis at the expense of growth, storage, defense, and reproduction [73]. Fitness of an individual organism can be measured in terms of Scope for Growth (SfG), which is the measurement of physiological energy balance and it ranges from optimal (positive values) to stressed conditions (negative values) when the organism is exposed to contaminants or unfavorable environmental conditions [74, 75]. The SFG has been widely used in field monitoring studies [76, 77]. The SFG and the growth rates of mussels were drastically reduced when mussels from uncontaminated sites were transplanted A wide variety of chemical contaminants capable of directly or indirectly damaging the DNA of marine organisms are being discharged into the marine environment. These genotoxic chemicals are capable of inducing some changes in the molecular and cellular levels of marine bivalves [91, 92]. Two well-known tests, micronucleus assay and comet assay, are being widely used to assess the genotoxic effects of environmental contaminants on marine bivalves [91, 92]. The micronucleus assay is used to detect the structural and numerical chromosomal changes while the comet assay (single-cell gel electrophoresis) is used to detect DNA strand breaks in marine bivalves.

## **6. Coastal pollution monitoring using biomarkers a case study**

The biomarkers in marine bivalves based on sub-lethal effects of contaminants are ecologically relevant and can be used to give subtle signals of response to contaminants before damage becomes irreversible. The water quality in European coastal sites was classified ranging from class 1 (clean areas) to class 5 (highly polluted areas), based on global biomarker index for Baltic mussels [93]. The Marine Strategy Framework Directive (Directive 2008/56/EC) since 2008 emphasized on the importance of assessing key biological responses for evaluating the health of organisms and linking the observed changes to potential contaminant effects [94].

The studies conducted prior to 1990s from Puget Sound, Washington, reported high concentrations of toxic metals, polycyclic aromatic hydrocarbons (PAHs) and PCBs in sediments and toxicant-induced, adverse effects in benthic fish samples collected from the urban associated sites [95]. As an example of how biomarker-based indices can be integrated into environmental monitoring of Puget Sound, biomonitoring study using mussels was conducted in 1992 [7]. Blue mussels (*Mytilus edulis*) were collected from their natural beds from nine sites in Puget Sound (**Figure 2**). Sites included the minimally contaminated reference areas of Oak Bay, Coupeville, and Double Bluff, in central and north Puget Sound, and Saltwater Park of south Puget Sound. Urban sites that were sampled for mussels included Eagle Harbor, Seacrest and Four Mile Rock in Elliott Bay, City Waterway in Commencement Bay, and Sinclair Inlet.

Relatively high tissue concentrations of contaminants including toxic trace metals were observed in mussels tissue samples from the urban-associated sites compared to the minimally

> contaminated (reference) sites (**Figure 3**). Mussels from contaminated sites showed low LMS, enhanced lipofuscin deposition, and increased accumulation of lysosomal and cytoplasmic unsaturated neutral lipids (**Figure 3**). Mussels from the contaminated sites were smaller in size together with lower somatic tissue weight relative to shell length [7]. Highly significant correlations were observed between tissue concentrations of selected toxic elements (measures of anthropogenic exposure) and LMS [7]. The study showed that biomarkers in mussels have the potential to be used as sensitive, accurate, and rapid techniques for assessing the biological impact of environmental contaminants in the coastal waters. The study results were in agreement with the previous study results, which showed an association between metabolites of aromatic compounds in bile and the occurrence of hepatic lesions in English sole (*Parophrys*

> **Figure 3.** Relationship between lysosomal membrane stability (LMS) and tissue concentration of heavy metals (mercury

Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs

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165

and lead) of mussels from urban-associated and reference sites in Puget Sound [7].

Commercially and ecologically important marine bivalves (clams, mussels, and oysters) are widely used for monitoring levels of trace metals in the marine environment from several parts of the world. Trace metal monitoring using bivalves has several advantages compared to using seawater or sediment samples for the same purpose. Bivalves such as mussels are having global distribution from the polar to the tropical region and being successfully used for temporal and spatial trend monitoring of trace metals in the coastal waters across the globe. Recently several biomarkers, the biological responses of bivalves to contaminants including trace metals, are being developed and tested to assess the coastal water quality. The biomarkers of stress in bivalves give early warning signal about the presence of toxic trace metals in the marine environment.

*vetulus*) from Puget Sound [96].

**7. Conclusion**

**Figure 2.** Map showing the mussel sampling sites in Puget Sound, Washington [7].

**Figure 3.** Relationship between lysosomal membrane stability (LMS) and tissue concentration of heavy metals (mercury and lead) of mussels from urban-associated and reference sites in Puget Sound [7].

contaminated (reference) sites (**Figure 3**). Mussels from contaminated sites showed low LMS, enhanced lipofuscin deposition, and increased accumulation of lysosomal and cytoplasmic unsaturated neutral lipids (**Figure 3**). Mussels from the contaminated sites were smaller in size together with lower somatic tissue weight relative to shell length [7]. Highly significant correlations were observed between tissue concentrations of selected toxic elements (measures of anthropogenic exposure) and LMS [7]. The study showed that biomarkers in mussels have the potential to be used as sensitive, accurate, and rapid techniques for assessing the biological impact of environmental contaminants in the coastal waters. The study results were in agreement with the previous study results, which showed an association between metabolites of aromatic compounds in bile and the occurrence of hepatic lesions in English sole (*Parophrys vetulus*) from Puget Sound [96].

## **7. Conclusion**

Baltic mussels [93]. The Marine Strategy Framework Directive (Directive 2008/56/EC) since 2008 emphasized on the importance of assessing key biological responses for evaluating the health of organisms and linking the observed changes to potential contaminant effects [94]. The studies conducted prior to 1990s from Puget Sound, Washington, reported high concentrations of toxic metals, polycyclic aromatic hydrocarbons (PAHs) and PCBs in sediments and toxicant-induced, adverse effects in benthic fish samples collected from the urban associated sites [95]. As an example of how biomarker-based indices can be integrated into environmental monitoring of Puget Sound, biomonitoring study using mussels was conducted in 1992 [7]. Blue mussels (*Mytilus edulis*) were collected from their natural beds from nine sites in Puget Sound (**Figure 2**). Sites included the minimally contaminated reference areas of Oak Bay, Coupeville, and Double Bluff, in central and north Puget Sound, and Saltwater Park of south Puget Sound. Urban sites that were sampled for mussels included Eagle Harbor, Seacrest and Four Mile Rock in Elliott Bay, City Waterway in Commencement Bay, and Sinclair Inlet.

164 Trace Elements - Human Health and Environment

Relatively high tissue concentrations of contaminants including toxic trace metals were observed in mussels tissue samples from the urban-associated sites compared to the minimally

**Figure 2.** Map showing the mussel sampling sites in Puget Sound, Washington [7].

Commercially and ecologically important marine bivalves (clams, mussels, and oysters) are widely used for monitoring levels of trace metals in the marine environment from several parts of the world. Trace metal monitoring using bivalves has several advantages compared to using seawater or sediment samples for the same purpose. Bivalves such as mussels are having global distribution from the polar to the tropical region and being successfully used for temporal and spatial trend monitoring of trace metals in the coastal waters across the globe. Recently several biomarkers, the biological responses of bivalves to contaminants including trace metals, are being developed and tested to assess the coastal water quality. The biomarkers of stress in bivalves give early warning signal about the presence of toxic trace metals in the marine environment.

## **Acknowledgements**

We thank the Center for Environment and Water, Research Institute, King Fahd University of Petroleum and Minerals, Dhahran, Saudi Arabia, for providing research facilities. We acknowledge the research funding (# T.K. 11-0629) of the King Abdulaziz City for Science and Technology (KACST). We also thank our colleagues and students who helped us to prepare this manuscript.

[7] Krishnakumar PK, Casillas E, Varanasi U. Effects of environmental contaminants on the health of *Mytilus edulis* from Puget Sound, Washington, USA. I. Cytochemical measures of lysosomal responses in the digestive cells using automatic image analysis. Marine

Biomonitoring of Trace Metals in the Coastal Waters Using Bivalve Molluscs

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## **Author details**

Periyadan K. Krishnakumar1 \*, Mohammad A. Qurban1 and Geetha Sasikumar<sup>2</sup>

\*Address all correspondence to: kkumarpk@kfupm.edu.sa

1 Center for Environment and Water, Research Institute, King Fahd University of Petroleum and Minerals, Dhahran, Saudi Arabia

2 Mangalore Research Centre, Central Marine Fisheries Research Institute, Mangalore, Karnataka, India

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## *Edited by Hosam El-Din M. Saleh and Eithar El-Adham*

Over the last few years, we have witnessed increasing efforts dedicated to the scientific investigation and characteristics of trace elements. Especially in the field of human and animal nutrition, trace elements display a considerably attractive issue for research because they play an essential role in the nutrition of both animals and humans. Aquatic environments contaminated with trace elements are an emerging research area due to the toxicity, abundance, and environmental persistence of trace elements. Accumulation of heavy metals as a class of trace elements in various environments, and the subsequent transition of these elements into the food and feed chain, severely affects human health.

The determination of type and concentration of trace elements is regarded as the first and most important step to follow the mechanisms controlling the dispersal and accumulation of trace elements. Element speciation in different media (water, soil, food, plants, coal, biological matter, food, and fodder) is pivotal to assess an element's toxicity, bioavailability, environmental mobility, and biogeochemical performance. Recently, new analytical techniques have been developed, which greatly simplified the quantitation of many trace elements and considerably extended their detection range. In this context, the development of reproducible and accurate techniques for trace element analysis in different media using spectroscopic instrumentation is continuously updated.

Published in London, UK © 2018 IntechOpen © Dmitr1ch / iStock

Trace Elements - Human Health and Environment

Trace Elements

Human Health and Environment

*Edited by Hosam El-Din M. Saleh* 

*and Eithar El-Adham*