Distributions of trace metals Co, Cu and Cd in northern Sagami Bay, Japan and their relationship to estuarine variables

Estuarine, Coastal and Shelf Science 111 (2012) 84e94

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Distributions of trace metals Co, Cu and Cd in northern Sagami Bay, Japan and their relationship to estuarine variables Hyoe Takata*,1, Tatsuo Aono, Shigeo Uchida National Institute of Radiological Sciences, 4-9-1 Anagawa Inage-ku, Chiba, 263-8555, Japan

a r t i c l e i n f o

a b s t r a c t

Article history: Received 1 December 2011 Accepted 26 June 2012 Available online 10 July 2012

Concentrations of the trace metals Co, Cu and Cd in dissolved (<0.2 mm) and particulate (>0.2 mm) phases in surface waters were determined to test whether DOC, which is an indicator of dissolved organic matter (DOM), and organic matter (OM) in suspended particulate matter (SPM) act as important estuarine variables controlling the distributions of these metals in the estuarine zone. Although the dissolved metals in the surface waters from the Sagami River to northern Sagami Bay had different behaviors in June, August, and November, 2008, concentrations of dissolved Co and Cu linearly correlated with DOC concentration, but with different slopes for fresh (salinity of
0.1) and estuarine (salinity of >0.1) waters. This difference in the water regime between those two metals and the DOC in Sagami Bay indicates that there are differences in composition between riverine and estuarine DOM, due to change of the composition by biogeochemical processes and the presence of additional sources, or that seawater cations are competing for binding sites of DOM during the estuarine mixing. As for particulate phase, there was no relationship between the concentrations of particulate metals and that of OM in SPM at low river flows (i.e., June and November surveys). It is thought that change of the OM composition by the additional SPM sources (anthropogenic discharge and resuspension from the seabed) contributes to the affinity of trace metals for sorption sites on the surfaces of the SPM, thereby making the evaluation of importance of OM in SPM difficult. On the other hand, at high river flow (August survey), riverine SPM significantly contributed to the estuarine SPM and it entered northern Sagami Bay with negligible composition change; however, there were weak correlation coefficients of particulate Co and Cd to OM in SPM, and there was no linear relationship between particulate Cu and OM in SPM. This can be attributed to the imbalanced competition between particle sorption and organic/inorganic complexation for Co and Cd, and prevention of particle sorption by binding Cu more strongly to dissolved organic ligands. Ó 2012 Elsevier Ltd. All rights reserved.

Keywords: Trace metals Dissolved organic carbon Organic matter in suspended particulate matter Salinity Distribution coefficients

1. Introduction Trace metals in rivers, which are distributed between particulate (>0.2e0.45 mm fraction) and dissolved (<0.2e0.45 mm fraction) phases, are transported to the oceans. In estuaries, which encompass the river and ocean interface leading to axial salinity gradients, the distributions of estuarine variables (e.g., salinity, alkalinity and pH) are subject to rapid change. The difference of water conditions in freshwater and estuarine seawater can cause a redistribution of metals between particulate and dissolved phases. Suspended particulate matter (SPM) plays an important role in controlling the behavior and fate of trace metals because the

* Corresponding author. E-mail addresses: [email protected], [email protected] (H. Takata). 1 Present address: Marine Ecology Research Institute, 347 Yamabuki-cho, Shinjuku-ku, Tokyo, 162-0801, Japan. 0272-7714/$ e see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.ecss.2012.06.009

removal or sorption of trace metals by SPM in aquatic environments regulates the concentration, speciation and bioavailability of trace metals in estuarine waters (Baskaran and Santschi, 1993; Turner, 1996; Turner and Millward, 2000, 2002; Martino et al., 2002). The distribution of carrier-phase metals (i.e., Al, Mn and Fe) in estuaries is dominated by particle dynamics (Tang et al., 2002); however, that of other trace metals is strongly influenced by the dynamics of organic matter (OM) cycling and the production of organic ligands of different sizes (Sholkovitz, 1976; Santschi et al., 1997). A number of studies have reported the importance of OM in controlling concentrations of trace metals in estuarine and coastal waters (Adediran and Kramer, 1987; van den Berg et al., 1987; 1990; Santschi et al., 1997; Baeyens et al., 1998a; Wen et al., 2008). Distributions and fluxes of OM in SPM in estuaries appear to be regulated by the mixing and different settling of a complex set of source materials (e.g., eroded coastal material, atmospheric particles, biogenic matter formed in situ, and contaminated particles

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

derived from local dumping of industrial and sewage solids) during transit through the estuarine mixing zone (Thornton and McManus, 1994; Millward et al., 1996; Middelburg and Nieuwenhuize, 1998; Middelburg and Herman, 2007). Furthermore, OM composition is probably changed by biogeochemical reactivity (Middelburg and Herman, 2007). For example, incorporation of nitrifiers and heterotrophic bacteria (Caraco et al., 1998; Middelburg and Nieuwenhuize, 2000a,b) may elevate bulk d15N values in organic matter. The additional sources and change of composition of OM during the estuarine mixing affect abundance or availability of sorption sites. For trace metals which have a sufficient affinity for dissolved organic matter (DOM), studies have shown that organic complexation in the dissolved phase is important for trace metals in estuarine waters (Zhang et al., 1990; Santschi et al., 1997; Baeyens et al., 1998b; Muller, 1999; Bruland et al., 2000; Buck and Bruland, 2005; Santos-Echeandía et al., 2008). Although riverine humic or fulvic acids are responsible for the organic complexation in estuarine waters (Kogut and Voelker, 2001; Hamilton-Taylor et al., 2002), industrial and sewage discharges might influence the composition of dissolved organic ligands in estuaries: EDTA-like ligands and activated sludge biopolymers as anthropogenic ligands are known to be responsible for complexation of metals (Sedlak et al., 1997; Bedsworth and Sedlak, 1999). Phytoplankton can excrete highaffinity, hydrophilic metal-specific complexing organic ligands, as well as colloids with metal-complexing functionality (Muller, 1999). Metals strongly binding to dissolved organic ligands, additionally, are thought to be protected from the seawater ion competition caused by changes in salinity in the estuarine zone (Hamilton-Taylor et al., 2002). Thus, the size distribution of both OM and trace metals can exhibit the competitive effects of particle sorption and complexation of metals with dissolved organic ligands (or inorganic species), and this distribution is important for understanding the particleewater interactions of trace metals in estuaries. In a previous study, we sought sources of Co in northern Sagami Bay which is connected to the Sagami River, and the results suggested that benthic remobilization, desorption from SPM, and sewage farms could be important sources for it in the estuarine area (Takata et al., 2010). In the present study, we further discuss the importance of dissolved organic carbon (DOC, as an indicator of DOM) and OM in SPM in controlling behavior of this metal in the estuarine waters collected in northern Sagami Bay. Distributions of Cu and Cd in northern Sagami Bay are also presented. Understanding of particleewater interactions of those two metals in estuarine areas is very necessary because Cu and Cd are highly toxic metals that affect biological activities in marine environments (Xue and Sunda, 1997; Hudson, 1998; Porter et al., 2004). As well, we investigate the sources and behavior of DOC and OM in Sagami Bay and then test whether DOC and OM in SPM as well as other estuarine variables (e.g., salinity, pH and redox conditions) play an important role in affecting behavior of the trace metals during the estuarine mixing. Data on salinity, pH, and concentrations of SPM, dissolved oxygen (DO) and DOC in the estuarine and coastal area are cited from our earlier work (Takata et al., 2010). 2. Methods 2.1. Sampling Sampling locations are shown in Fig. 1. The entire area of Sagami Bay is about 25,002 km2. The northern section is about 36 km2 in area with a mean depth of 30 m. Samples were collected along a salinity gradient across the Sagami River and northern Sagami Bay in three different surveys in 2008 in which the mean river water

85

discharges were 39 m3/s (June), 139 m3/s (August), and 20 m3/s (November). Apparent residence times of freshwater, which were calculated as the freshwater volumes (estimated using the salinity distribution) divided by the discharges, were approximately 3 d (June), 11 d (August), and 2 d (November) for northern Sagami Bay. The sampling method and sample treatment were described previously (Takata et al., 2010). In brief, one or two surface water samples at each station were collected from northern Sagami Bay using an acid-cleaned, Teflon-coated 5 L horizontal Niskin X water sampler (General Oceanics) deployed from a boat. The samples for the dissolved trace metals (<0.2 mm fraction) were gravity filtered into precleaned low-density polyethylene (LDPE) bottles by connecting a precleaned 0.2 mm pore size polytetrafluoroethylene filter (capsule cartridge type, Advantec) to each sampling bottle spigot. Unfiltered water samples for determining total dissolvable metals were placed in precleaned LDPE bottles. The filtered and unfiltered water samples for metal analyses were acidified to pH w1.5 with 15.3 M nitric acid solution. The rest of the water in the sampler was used for the measurements of SPM, nutrient, particulate organic carbon (POC), particulate organic nitrogen (PON), DOC, and DO concentrations. POC, PON and DOC samples were collected in precombusted amber glass bottles, and immediately placed in an icebox and brought back to the laboratory where we filtered them through pre-combusted (450  C) Whatman GF/F filters (pore size 0.7 mm). The filters for POC and PON analyses were then air-dried. The volumes of the filtrates were measured for concentration determination. The filters and filtrates for DOC measurement were stored in a refrigerator until analysis. 2.2. Analysis In our previous study (Takata et al., 2010), we analyzed acidified water samples without UV digestion, which is commonly used for photo-oxidative degradation of organic matter in water samples. In the present study, the nitric acid-acidified filtered samples were added into quartz glass flasks and irradiated for 24 h with a 100-W high-pressure Hg-vapor UV lamp (OHD-110M-ST, ORC Manufacturing Co., Ltd.). Then, these UV-irradiated filtered samples were stored in the dark until the metal analyses. We determined total dissolvable metal concentrations using the acidified unfiltered water samples. The acidified unfiltered water samples, which were stored for about two months, were filtered through precleaned 0.2 mm pore size polycarbonate filters (Advantec), and then, the water samples were irradiated with the UV lamp. The total dissolvable fraction included surface desorbed particulate and dissolved metals. For determination of the particulate metal, concentrations of total dissolvable metal were subtracted from those of total dissolvable metal. The obtained concentrations of particulate metal (w/v) were converted to w/w using SPM concentrations. Generally, acetate buffer, acetic acid or methanesulfonic acid solutions are used to extract surface adsorbed metals on particles retained on filters (Lewis and Landing, 1992; Tang et al., 2002; Wen et al., 2008). However, we used nitric acid to dissolve metals in the unfiltered water samples. Nitric acid is a stronger acid than acetic acid or methanesulfonic acid and will extract more metals than the weaker acids at equal concentrations. In the present study, however, the pH of the acidified unfiltered water samples was w1.5 and similar to that of an acetic acid solution (about 1.9, Lewis and Landing, 1992) or a methanesulfonic acid solution (about 1.7, Tang et al., 2002). Significant amounts of the trace metals that were extracted are assumed to be weakly bound in an exchangeable fraction associated with particle surfaces. Thus, most of the particulate metal defined in this study was considered to be adsorbed on the surface of the suspended particles.

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H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

Fig. 1. Sampling sites in northern Sagami Bay, Japan. Stations were sampled in June, August and November 2008.

An inductively coupled plasma mass spectrometer equipped with an octapole collision reaction cell (Agilent 7500c, Yokogawa Analytical Systems), as described in the literature (Takata et al., 2009), was used in analyses for the metals Co, Cu and Cd in water samples that had been treated by solid-phase extraction with NOBIAS-CHELATE PA1 chelating resins (polyamines and polycarboxyl functional groups immobilized on hydrophilic methacrylates; Hitachi High-Tech) (Sakamoto et al., 2006). The accuracy of the analytical procedure was checked using a reference material, estuarine water, SLEW-3, and near-shore seawater, CASS-4 (National Research Council of Canada) (Table 1). Our concentrations showed close agreement with the certified values. The precision was estimated to be better than 10% at the trace metal concentrations in the reference materials. Detection quantification values from analyzing de-ionized water were 0.16, 2.4 and 0.1 ng/L for Co, Cu and Cd, respectively. Table 1 Analytical results (n ¼ 5) for the certified reference materials of estuarine water SLEW-3 and near-shore seawater CASS-4. Element

SLEW-3 (mg/L) Co Cu Cd CASS-4 (mg/L) Co Cu Cd

Certified value

Measured value

Average  SD

Average  SD

0.042  0.01 1.550  0.120 0.048  0.004

0.042  0.001 1.487  0.032 0.049  0.003

0.026  0.003 0.592  0.055 0.026  0.003

0.025  0.001 0.548  0.004 0.026  0.001

For POC and PON measurements, after weighing the air-dried filters, we slowly added 1 M HC1 to remove inorganic carbon from the samples. The filters were dried again and placed into tin cups. The individual cups were then pinched, closed, and compacted into a ball. These balls were used for determination of POC and PON concentrations with a Perkin-Elmer CHN analyzer (PE CHNS/O 2400). Average blank values were less than the detection limit (0.2 mgN) and 1.67  0.71 mg C, respectively. On the basis of these data we estimated the detection limit (three times the standard deviation of the blank) was about 2 mg C. The detection limit for nitrogen (about 0.2 mg N) was not determined by the blank, but by the sensitivity of the detector. The analytical precision of C and N analyses was less than 1% for a standard sample (acetanilide). Analyses of salinity, pH, DO, nutrient, SPM and DOC in the water samples were described in our previous study (Takata et al., 2010). 3. Results and discussion 3.1. Distributions of environmental variables as a function of salinity 3.1.1. DO and nutrient concentrations and pH DO concentrations were 8.5e10.2 mg/L (June), 7.1e9.0 mg/L (August), and 7.3e11.1 mg/L (November), suggesting oxic conditions existed in the surface waters in this study. The distributions of nutrient concentrations for each survey are shown in Fig. 2. Silicate showed a decrease in concentration with increasing salinity in the estuarine zone (Fig. 2A). Silicate concentrations in the low salinity zone (<10) ranged between 240 mM and 388 mM and then

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

decreased to less than 70 mM in the high salinity zone (>30). On the other hand, phosphate showed a non-conservative distribution during each survey (Fig. 2B). Sewage effluents might also influence high phosphate concentrations in the estuarine area (Yamada and Matsushita, 2006). There were no remarkable differences in the pH values (range: 7.4e8.4) within the water samples from each sampling station. 3.1.2. DOC Slight differences in DOC concentrations were observed among the three sampling surveys (Fig. 2C). The present DOC concentration range (0.6e1.4 mg/L) was relatively narrow compared to those in published reports (e.g., 0.8e2.4 mgC/L in San Francisco Bay ~ do-Wilhelmy et al., 1996); 4.6e5.9 mgC/L in the Lena River (Sanu Estuary and the Laptev Sea (Russia) (Garnier et al., 1996); 1.6e9.2 mgC/L in the Mersey Estuary (Martino et al., 2004)). The DOC levels were consistent with those in other Japanese estuarine areas (e.g., 0.5e2.0 mgC/L in Hiroshima Bay (Fukushima et al., 2001); 0.7e1.8 mgC/L in Gokasho Bay (Sakami, 2008); 0.5e1.5 mgC/L in southwestern Wakasa Bay (Takata et al., 2010)). DOC concentrations were slightly higher in estuarine water samples (salinity of >0.1) than in river water samples (salinity of
0.1) (e.g., the ratios of mean riverine DOC to mean estuarine DOC were 0.71 for June, 0.92 for August, and 0.65 for November). This observation was similar to the result reported for Hiroshima Bay by Fukushima et al. (2001). They revealed that an increase in DOC concentration resulted from autochthonous production of DOM within the bay, and chemical characteristics of DOC in the bay differed from those of DOC with river origins. 3.1.3. SPM, POC, PON, and OM in SPM concentration and C/N ratios Although concentrations of SPM during the August survey decreased slightly with increasing salinity, an increase in SPM concentration was observed within a salinity of 5e30 during June and November surveys. In addition, the relatively high SPM concentrations observed at northern Sagami Bay in August

B

June August November

300 200 100 0

POC (mgC/L)

E

D SPM (mg/L)

2.0 1.5 1.0 0.5 0.0

F

2.0

PON (mgN/L)

DOC (mgC/L)

C

(Fig. 2D), compared to the other surveys, were attributed to well developed turbidity resulting from the release of impounded water from the Miyagase dam upstream on the Sagami River. Both POC and PON displayed different behaviors during each survey (Fig. 2E,F). In August, they were mostly characterized by low salinity (
0.1) maxima, and by subsequent decreases with increasing salinity. In June and November, relatively low POC and PON concentrations were observed and those concentrations were constant along the salinity gradient, except for three isolated values at Stns. 10e12 in June (POC: 1.00e1.21 mgC/L; PON: 0.22e0.25 mgN/L). The exceptions would be due to the external input (e.g., resuspension from sediments and sewage particles). We discuss this in the next paragraph. Organic C/N molar ratios have been employed as source indicators of OM in SPM in natural waters (Thornton and McManus, 1994; Hedges et al., 1997; Middelburg and Nieuwenhuize, 1998; Herman and Heip, 1999; Middelburg and Herman, 2007). The C/N ratios in the June survey rapidly decreased from 9.4 to 5.8 at a low salinity range of 0.1e6.7, and then had constant values of 5.4e6.3 at salinity of 6.7e33.7 (Fig. 3A). In the November survey, the C/N ratios were nearly constant values of 4.8e5.8 at salinity of 0.1e31.7. However, the C/N ratios decreased in a narrow salinity range of 31.7e33.9. Decreases in C/N ratios in both June and November surveys were consistent with maximum values of phosphate (Fig. 2B). Interestingly, these results were observed at Stn. 3 in June and at Stn. 5 in November and both were located the near the mouth of the river. The rapid increase in phosphate was probably due to sewage input (Yamada and Matsushita, 2006). SPM showed variation in OM concentration at salinity >25, and in November, the concentration dramatically increased at salinity >31.9. The similarity in phosphate and C/N behaviors near the mouth of the river inferred that sewage input would indicate the addition of sewage solids by presenting different C/N ratios from riverine SPM. In June, additionally, slight increases in SPM concentration (Fig. 2D) were also observed between Stns. 2 and 5 near the mouth with shallow water depth (<5 m) where the flow of the river and tide (tidal

Phosphate (µM)

Silicate (µM)

A 400

1.5 1.0 0.5 0.0

0

5

10

15

20

Salinity

25

30

35

87

4 3 2 1 0 50 40 30 20 10 4 3 2 1 0 0.5 0.4 0.3 0.2 0.1 0.0

0

5

10

15

20

25

30

35

Salinity

Fig. 2. Concentrations of (A) silicate, (B) phosphate, (C) DOC, (D) SPM, (E) POC and (F) PON as a function of salinity in northern Sagami Bay. Data on silicate, phosphate, DOC, and SPM were cited from Takata et al. (2010).

88

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

A

the concentration of OM in SPM was nearly constant and was different from June and November survey concentrations. In consequence, sources of SPM in northern Sagami Bay had significantly differing characteristics between high river flow (August) and low river flow (June and November). The existence of additional sources of SPM (e.g., seabed and sewage solids) from distinctive differences in C/N ratio, behavior of SPM and phosphate, might affect the composition of the estuarine SPM under conditions of low river flows in June and November. At the high river flow in August, the large stock of riverine SPM might mask any SPM inputs.

12

C/N ratio

9 6 June August November

3 0

OM in SPM (µg/mg)

B 800 600 400 June November 200 August 0

0

5

10

15

20

25

30

35

Salinity Fig. 3. (A) C/N ratios and (B) concentration of OM in SPM as a function of salinity in northern Sagami Bay. Fitted lines are best-fit, exponential regression between the two parameters.

range: w2 m) form a high turbidity zone, resulting in the addition of sedimentary SPM from the seabed. Moreover, rapid increases in concentrations of POC, PON, and OM (POC þ PON) in SPM were observed at salinity >25 during June (Figs. 2E,F and 3B), indicating additional sources of OM in SPM; however, there was no significant difference in C/N ratios at salinity >25 in June. The distinction between sedimentary and riverine OM would fade away upon diagenesis (Hedges et al., 1997; Herman and Heip, 1999). Although relationships between particulate D13C, D15N and C/N ratios may allow the distinction of sources (Thornton and McManus, 1994; Middelburg and Nieuwenhuize, 1998; Middelburg and Herman, 2007), different sources of OM in SPM in northern Sagami Bay might be expected from the similarity in the phosphate, SPM, POC and PON, and C/N ratio behaviors during the June and November surveys. On the contrary, in the August survey, C/N ratios varied only slightly within 8.8e9.9 throughout the salinity gradient, and appeared to be independent of salinity changes. The mean C/N ratio (9.2  0.3, n ¼ 8) in estuarine water (salinity: >0.1) was comparable to that in freshwater (9.8  0.2, n ¼ 3). The mean ratio was consistent with that in the world’s rivers (C/N ratio: 8e10, Meybeck, 1982). Furthermore, studies on sources of suspended particles in estuarine and coastal regions have reported that riverine particles originate from degraded terrestrial plants and soils, resulting in high C/N ratios (Middelburg and Nieuwenhuize, 1998; Middelburg and Herman, 2007). Elevated phosphate concentrations were observed at Stns. 3 and 4 in August (Fig. 2B), indicating the potential source of sewage particles; however, the influence of this input might reflect loss of the source signature by incorporation with an overwhelming volume of riverine SPM within which it was thoroughly mixed and dispersed (Fig. 2D). We assumed that riverine SPM was probably responsible for the estuarine SPM at high flow in August. This assumption could also be explained by the results on concentration of OM in SPM, as depicted in Fig. 3B. During August,

3.1.4. Trace metals Water discharge from the Sagami River influenced the concentrations of Co, Cu and Cd in northern Sagami Bay. For example, the order of mean total dissolvable metals (particulate þ dissolved) in the estuarine waters was August (Co: 452 ng/L; Cu: 3.3 mg/L; Cd: 16 ng/L) > June (Co: 49 ng/L; Cu: 0.8 mg/L; Cd: 12 ng/L) > November (Co: 31 ng/L; Cu: 0.8 mg/L; Cd: 10 ng/L). The order was the same as that of river discharge, which was August (139 m3/s) > June (39 m3/ s) > November (20 m3/s). However, behavior of trace metals in the estuarine waters was independent of the hydraulic regimes. Although concentration of dissolved Co at Stns. 1 and 2 in the Sagami River (salinity of
0.1) varied in a narrow range in different river discharge months, dissolved Co had a low- or mid-salinity maximum during the three sampling surveys (Fig. 4AeC). Dissolved Cu concentration followed an approximate dilution line from the river to the seawater endmembers, although most of the data were so noisy that they did not provide proof of conservative behavior (Fig. 4DeF). Dissolved Cd showed seasonal variations in its behavior along the salinity gradient (Fig. 4GeI). During June and November, dissolved Cd exhibited a mid-salinity maximum in concentrations (Fig. 4G and I) as has been observed in other estuaries (Elbaz-Poulichet et al., 1987; Chiffoleau et al., 1994). In August, however, dissolved Cd concentration appeared to increase with increasing salinity (Fig. 4H). A mid-salinity maximum should exist in this month as well as in the June and November surveys and a slight mid-salinity increase seemed to appear. We thought that the mid-salinity increase was covered by the effect of the high flow rate in summer. Particulate metals in the Sagami River had seasonal variations in their concentrations (Fig. 4). Particulate Co and Cu in northern Sagami Bay in June remained roughly constant throughout the salinity gradient. In August, particulate Co and Cu showed a slight decrease along the salinity gradient. Particulate Cd data did not show salinity dependence during June and August. In November, concentrations of Co, Cu and Cd in the particulate phase at salinity <30 were relatively high compared to those at a salinity >30. 3.2. Distribution coefficients of metals between dissolved and particulate phases In order to describe the particleewater interactions of trace metals, we used the distribution coefficient, Kd (L/kg) (e.g., ~ do-Wilhelmy et al., 1996). The approChiffoleau et al., 1994; Sanu priate Kd values defining the particleewater interactivity of metals have been derived experimentally or empirically (Turner and Millward, 2002; Martino et al., 2002; Tang et al., 2002). The Kd of a particular metal between the dissolved and particulate phases is defined as the ratio of particulate metal concentration, [Mpar], to dissolved metal concentration, [Mdis], as given in Eq. (1).

  Kd ¼ Mpar ½Mdis 

(1)

Ranges of log Kd values of Co, Cu and Cd for northern Sagami Bay during the three sampling surveys are presented in Table 2. The Kd

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

June 2008

B

50

C

40 30 20 10

1.2

D

E

0 0.4

F

0.3

0.9 0.2 0.6 0.3

0.1

0.0 30

0.0 3

H

I

2

10

1

0 0

5

10 15 20 25 30 35 0

5

10 15 20 25 30 35 0

Salinity

Salinity

5

P-Cd (ng/mg)

G

20

P-Cu (µg/mg)

D-Cu (µg/L)

A

50

0 1.5

D-Cd (ng/L)

November 2008

August 2008

P-Co (ng/mg)

D-Co (ng/L)

100

89

0 10 15 20 25 30 35

Salinity

Fig. 4. Concentrations, as a function of salinity in northern Sagami Bay, of: (A-C) dissolved Co (D-Co) and particulate Co (P-Co); (D-F) dissolved Cu (D-Cu) and particulate Cu (P-Cu); and (G-I) dissolved Cd (D-Cd) and particulate Cd (P-Cd). The open circles and filled triangles represent dissolved and particulate metals, respectively.

for trace metals between particulate and dissolved phases had the order: Co (mean log Kd: 5.5) > Cu (5.2) > Cd (5.0). The trend for decreasing log Kd values among those metals was nearly the same as that previously reported (Valenta et al., 1986; Balls, 1989; Chiffoleau et al., 1994; Munksgaard and Parry, 2001) (Table 2). There were wide variations in log Kd values for the metals in each sampling month. Except for log Kd values for Cu in August and November, log Kd values for all the metals varied by approximately one order of magnitude regarding their differences between minimum and maximum values. Salinity affects the partitioning of trace metals through the competitive, complexing and electrostrictive effects of seawater ions in estuarine and coastal waters (Gschwend and Schwarzenbach, 1992; Turner and Millward, 2002). For several metals, the Kd values are reduced by one or two orders of magnitude by an increase in salinity moving from a river to the sea (Turner and Millward, 2002 and references therein) because of competitive adsorption by seawater cations and the formation of stable and soluble chloro-, sulfato-, carbonato, and organiccomplexes in seawater.

Table 2 Log Kd (L/kg) for trace metals in northern Sagami Bay and other locations. Estuary

Co

Cu

Cd

References

Seine estuary Scheldt estuary British coast North Australian

5.0

4.5e4.7 4.8e5.0 4.0e5.0 3.7e5.4

3.9e4.0 4.5e5.0 3.5e5.0 3.3e6.3

Chiffoleau et al. (1994) Valenta et al. (1986) Balls (1989) Munksgaard and Parry (2001)

3.0e5.7 4.5e5.9

3.8e6.2

3.1e3.9 4.6e5.7

Garnier et al. (1996) This study

5.4e6.2 4.4e6.4

5.0e5.4 5.3e5.7

4.3e5.7 4.2e5.9

Coast and estuaries Lena River estuary Sagami Bay (June 2008) (August 2008) (November 2008)

Plots of the Kd values of metals as a logarithm scale in northern Sagami Bay vs. salinity are shown in Fig. 5A,C,E. Decrease in log Kd along the salinity gradient was observed for Co during the three surveys (r < 0.4, p < 0.1). This trend was observed for Cd in August and November (r < 0.8, p < 0.1), while no remarkable salinitydependent decrease in log Kd of Cd was observed in June. The decline for those two metals suggested that salinity affected particleewater interactions of metals, consistent with other studies (Turner, 1996; Garnier et al., 1996; Tovar-Sánchez et al., 2004). For Cu, Kd values were scattered or constant throughout the salinity gradient, probably indicating that the Kd values of Cu were independent of salinity change in northern Sagami Bay. Previous studies have demonstrated inverse linear relationships between metal log Kd and SPM concentration when SPM concentration is plotted on a log scale (Benoit et al., 1994; Tang et al., 2002 and references therein). These relationships are explained by the particle concentration effect which is attributed to an increased colloid concentration in the filtered fraction in proportion to the quantity of SPM retained by the filter (Honeyman and Santschi, 1988, 1989; Tang et al., 2002). The inverse linear log Kd vs. log SPM relationship for trace metals suggested operation of the particle concentration effect in estuarine waters. When the log Kd values for trace metals were plotted against log SPM (Fig. 5B,D,F), a particle concentration effect, i.e., an inverse dependency of Kd on SPM, was more pronounced for Cu in June and August (Fig. 5D), indicating the presence of colloidal species. In November, the effect was not observed for Cu. Since relatively low SPM concentrations were observed in November, compared to the other months, colloidally bound forms were not predominant in the filter-passing fraction (<0.2 mm) in November. For Co, although a positive correlation was observed in November, this metal could be dependent on salinity, rather than particle concentration (Takata et al., 2010). For Cd, there was no significant relationship between log Kd and log SPM in the three sampling months (Fig. 5F). It has been reported that most of the dissolved Cd (w100%) species were

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LogKd-Co

7

A

B

August November r = –0.90, p < 0.05 r = –0.51, p < 0.1

6

5

June r = –0.47, p < 0.05

4 7

D

C

LogKd-Cu

6

August r = –0.56, p < 0.01

5 June August November

4

June r = –0.61, p < 0.001

3 7

E

LogKd-Cd

6

F November r = –0.80, p < 0.05

5

4

August r = –0.80, p < 0.005

3 0

10

20

30

Salinity

40

-1

10

10

0

1

10

10

2

SPM (mg/L)

Fig. 5. Distribution coefficient, Kd, of metals versus salinity, and Kd of metals versus SPM in northern Sagami Bay.

found in the
1 kDa or 10 kDa size fractions not in the colloidal fraction (Kraepiel et al., 1997; Wen et al., 2006). 3.3. Relationship between trace metals and DOC The relationships between dissolved trace metal concentration and DOC concentration in northern Sagami Bay are plotted in Fig. 6A,C,E. No significant relationship appeared between dissolved Cd and DOC, indicating that dissolved Cd was mainly present as weak complexes or ionic pairs. This result was consistent with literature results that the degree of organic complexation was estimated to be less than 30% of total dissolved Cd in the Scheldt estuary (Baeyens et al., 1998b) and that the dissolved Cd was mostly present as cadmium-chloride complexes in the Gironde estuary (Kraepiel et al., 1997). Similar observations were obtained in our laboratory experiments (Takata et al., 2012) where the fraction of

organic Cd complexed was estimated to less than 30% under estuarine conditions. This result indicates that the effect of organic complexation of this metal on the particleewater interactions is relatively little. On the other hand, log Kd of Cd in the estuarine waters exhibited a reduction along the salinity gradient (Fig. 5E). The relationship is attributed to competitive adsorption by seawater cations and the formation of stable and soluble chloro-, sulphato- and carbonato-complexes in seawater (Paalman et al., 1994; Baeyens et al., 1998a,b; Turner et al., 2008). Dissolved Co and Cu concentrations tended to increase with increasing DOC concentration; however, it was apparent that the slopes for concentrations of dissolved Co and Cu vs. that of DOC (25.9 ng/L dissolved Co: mgC/L DOC; 1.08 mg/L dissolved Cu: mgC/L DOC) obtained in the Sagami River water samples (salinity of
0.1) differed from those (103 ng/L dissolved Co: mgC/L DOC; 1.53 mg/L dissolved Cu: mgC/L DOC) in estuarine water samples (salinity of

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

100

7

A

B

r= -0.61 p<0.2

r=0.89 p<0.05

25 0 1.5 1.2

D-Cu (µg/L)

LogKd-Co

50

C

6

5

4 6

r=0.98 p<0.005

LogKd-Cu

D-Co (ng/L)

75

r=0.75 p<0.01

0.9 0.6

D

5

r= -0.68 p<0.2

r=0.72 p<0.0001

0.3 0.0 15

4 7

F

E 10

6

LogKd-Cd

D-Cd (ng/L)

91

5

0 0.50

0.75

1.00

1.25

1.50

DOC (mg/L)

5

4 0.50

0.75

1.00

1.25

1.50

DOC (mg/L)

Fig. 6. (A) Dissolved Co (D-Co), (B) Kd-Co, (C) dissolved Cu (D-Cu), (D) Kd-Cu, and (E) dissolved Cd (D-Cd), and (F) Kd-Cd versus DOC for (filled squares) freshwater (salinity of
0.1) and (open circles) estuarine water (salinity of >0.1) in northern Sagami Bay.

>0.1). In addition, dissolved Co and Cu concentrations in the Sagami River water (salinity of
0.1) were significantly correlated with that of DOC (r > 0.8, p < 0.05 for Co and Cu). There was a weak correlation coefficient (r < 0.8) between both dissolved metals and DOC for the estuarine water samples. A similar relationship between Kd value and DOC concentration was observed for trace metals (Fig. 6B,D,F), although no significant relationship was observed in the estuarine water samples. The distinctive different relationships between Co and Cu and DOC in river water samples and between those two metals and DOC in estuarine water samples in northern Sagami Bay might reflect differences in the degree of organic complexation because of a change in composition of DOM. Recently, excitation emission matrix fluorescence and parallel factor analysis results were reported that DOM in the estuarine zone of Ise Bay, Japan was characterized by several different fluorescence components (riverine, terrestrial, and autochthonous origins), and change of fluorescence characteristics by estuarine biological and/or microbial activities was observed (Yamashita et al., 2008). Similar observations have been reported from a laboratory experiment (de Souza Sierra et al., 1997), in the Orinoco River which flows into the Caribbean Sea (Del Castillo et al., 1999), in two Maine estuaries (Mayer et al., 1999), and in a Danish estuary and its catchment (Stedmon and Markager, 2005). Therefore, we assumed that there were changes of

composition, additional sources, and autochthonous production for DOM in the estuarine zone in northern Sagami Bay. In addition, there is an increasing competition with Co for dissolved organic ligands from divalent seawater cations (e.g., Ca and Mg), and increasing complexation of metals with inorganic ligands as salinity increases (Hamilton-Taylor et al., 2002). In our laboratory experiments, speciation calculations predicted that although about 90% of dissolved Co was complexed by organic ligands in river water (salinity of £0.1), the fraction decreased to 60% in seawater (salinity of 31, Takata et al., 2012). Some previous studies also report that log Kd for Co shows a decline along the salinity gradient (Li et al., 1984; Garnier et al., 1996). The slight salinity-dependent reduction in organic complexation can be resulting in a balance between increasing cation competitions and increasing complexation with inorganic anions along the salinity gradient. As for Cu, there was no significant correlation between log Kd of this metal and DOM concentration. It is thought that the effect of DOM on particleewater interactions of this metal was not evaluated sufficiently as changes in concentration of DOC (0.6e1.4 mg/L) in this bay were less intense for other studies (e.g., 0.8e2.4 mgC/L in ~ do-Wilhelmy et al., 1996); 1.6e9.2 mgC/L in San Francisco Bay (Sanu the Mersey Estuary (Martino et al., 2004)). However, in our laboratory experiments revealed that the most of dissolved Cu was estimated to be organically complexed (mean percentage: 60%)

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

through the salinity gradient (Takata et al., 2012). This result indicates that protection of Cu from adsorption by binding this metal more strongly to the organic ligands. Moreover, the particle concentration effect was observed (Fig. 5D), that is, the presence of Cu associated with colloidal matter could be included in the dissolved phase (<0.2 mm fraction). When colloidal Cu increases with increasing SPM concentration, dissolved Cu data might overestimate the truly soluble Cu species.

6.5 Co r= 0.77, p < 0.05

6

Kd (L/kg)

92

5 Cd r= 0.61, p < 0.05

3.4. Relationship between OM in SPM and particulate trace metals Plots of particulate metals versus OM in SPM are shown in Fig. 7. No linear relationship for particulate metals was observed during June and November. This was not surprising as changes in composition of OM by the additional sources of SPM (anthropogenic discharge, and resuspension from the seabed) contributed to the affinity of trace metals to binding sites of OM in SPM. In June and November surveys, there were changes in C/N ratios of OM in SPM and rapid increases in phosphate during the estuarine mixing (Figs. 2B and 3A). These phenomena inferred the addition of sewage solids. Additionally, in June, increases in SPM, POC, and PON concentration (Fig. 2DeF) were also observed between Stns. 2 and 5 near the mouth, which forms the high turbidity zone, resulting in the addition of sedimentary SPM from the seabed. Previous studies argued that the relationship between metals and particles was dependent on particle compositions (Santschi et al., 1997; Wen et al., 2008). If the degrees of the affinity of trace metals were different among riverine, resuspended, and sewage particles, particle metals could result in nonlinear relationships with OM in SPM, probably thereby masking the importance of OM in SPM on the distributions of particulate metals at low river flow conditions.

P-Co (ng/mg)

A

50 August 40 r = 0.50, r < 0.1

June August November

30 20 10 0

P-Cu (µg/mg)

B

0.6 0.5 0.4 0.3 0.2 0.1 0.0

P-Cd (ng/L)

C

4

0

30

60

90

120

OM in SPM (µg/mg) Fig. 8. Distribution coefficients (Kds) of Co (open circles), Cu (filled triangles) and Cd (crosses) versus concentration of OM in SPM during the August survey.

In August, riverine SPM accounted for the estuarine SPM significantly and it was carried into northern Sagami Bay with negligible change of OM compositions. Although particulate Co and Cd correlated with OM in SPM in August, correlation coefficients were 0.50 (p < 0.1) for particulate Co and 0.73 (p < 0.05) for particulate Cd. As for Cu, it has been reported that this metal appears to be associated with OM in SPM (Adediran and Kramer, 1987; Baeyens et al., 1998a; Wen et al., 2008). In this study, particulate Cu did not exhibit a correlation with OM during August (Fig. 7B). The Kds of trace metals are plotted against OM in SPM during August in Fig. 8. If trace metals are sorbed on OM in SPM, Kds will exhibit a positive dependency on the concentration of OM in SPM. In northern Sagami Bay, however, Kd for Cu was independent of concentration of OM in SPM. Assuming that organic complexation regulated dissolved Cu speciation in northern Sagami Bay as shown in Fig. 6C, dissolved organic ligands could function to buffer metal ion activities in the estuarine waters, thereby decreasing or preventing metal sorption to particulates. Previously, laboratory experiment and model analysis indicated that Cu more strongly binds to dissolved organic ligands than Co and Cd under estuarine conditions (Hamilton-Taylor et al., 2002). Moreover, particle concentration effect was more pronounced for Cu at the high river flow in August (Fig. 5D), indicating the presence of colloidally bound forms in the filter-passing fraction (<0.2 mm). Thus, the independency would be due to the prevention of Cu particle sorption by binding Cu more strongly to the organic ligand or presenting its metal as the colloidal forms. As for Co and Cd, Kd had weak correlations with OM in SPM during the August sampling period (Fig. 8). This result implied that OM in SPM contributed to the affinity of trace metals for the sorption sites on the surfaces of the SPM. However, since the Kds of those two metals had salinity dependency (Fig. 5A,E), the observed weak correlations could also reflect the imbalance of a competition between particle sorption and organic/inorganic complexation along the salinity gradient.

4

August 3 r = 0.73, r < 0.005

4. Conclusions

2 1 0

0

100

200

300

400

500

600

700

OM in SPM (µg/mg) Fig. 7. Concentrations of (A) particulate Co (P-Co), (B) particulate Cu (P-Cu) and (C) particulate Cd (P-Cd) versus concentration of OM in SPM.

Concentrations of trace metals (Co, Cu and Cd) in dissolved (<0.2 mm fraction) and particulate (>0.2 mm fraction) phases for river, estuarine and coastal waters across a salinity gradient in northern Sagami Bay were used to investigate the fate and behavior of the metals. Dissolved and particulate metals showed variations in their behavior during three different surveys in June, August, and November 2008. We evaluated the importance of DOM and OM in SPM on particleewater interactions of Co, Cu and Cd in the estuarine waters collected in northern Sagami Bay, which connects to the Sagami River. Dissolved Co and Cu concentrations plotted against

H. Takata et al. / Estuarine, Coastal and Shelf Science 111 (2012) 84e94

DOC concentrations gave lines with different slopes between fresh (salinity of
0.1) and estuarine (salinity of >0.1) waters. This different water regime between those two metals and DOC in the bay was probably due to the changes in composition of DOM, the presence of additional sources, or seawater cation competition for binding sites of DOM during estuarine mixing. The behavior and sources of SPM in northern Sagami River had significantly differing characteristics between high river (August) and low river (June and November) flows. The existence of additional sources of SPM (e.g., seabed and sewage solids), which were seen from distinctive differences in C/N ratio and in behavior of SPM and phosphate, might affect the composition of the estuarine SPM under conditions of low river flows in June and November. At high river flow in August, the large stock of riverine SPM contributed significantly to the estuarine SPM; however weak correlation coefficients of particulate Co and Cd to concentration of OM in SPM were observed, and there was no linear relationship between particulate Cu and OM in SPM. As for Co and Cd, from results of relationships between distribution coefficients (Kds) for Co and Cd and salinity and between Co and DOM, the imbalance of a competition between particle sorption and organic/inorganic complexation during the salinity gradient could affect the particleewater interactions of those two metals. As for Cu in the estuarine zone, strongly binding this metal to dissolved organic ligands and/or presenting its metal as the colloidal forms was thought to prevent particle sorption on the metals.

Acknowledgments We are grateful for helpful comments on the manuscript from two anonymous reviewers. We thank S. Yamano, M. Matsui, M. Tomita, O. Hirokawa, Y. Endo, and R. Asada (KANSO Technos Co., Ltd.) for assistance during sampling and J. Shirasaka (Tokyo Nuclear Service) for general technical assistance. This work has been partially supported by the Agency for Natural Resources and Energy, and the Ministry of Economy, Trade and Industry (METI), Japan.

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