Impact of disposal of dredged material on sediment quality in the Kaohsiung Ocean Dredged Material Disposal Site, Taiwan

Chemosphere 191 (2018) 555e565

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Impact of disposal of dredged material on sediment quality in the Kaohsiung Ocean Dredged Material Disposal Site, Taiwan Chih-Feng Chen a, Chiu-Wen Chen a, Yun-Ru Ju a, Chih-Ming Kao b, Cheng-Di Dong a, * a b

Department of Marine Environmental Engineering, National Kaohsiung Marine University, Kaohsiung, Taiwan Institute of Environmental Engineering, National Sun Yat-Sen University, Kaohsiung, Taiwan

h i g h l i g h t s
The heavy metals pollution in core sediments from KODMDS was evaluated.
A slight increase of heavy metals concentration in surface sediments of KODMDS.
Dumping dredged sediment may raise potentially ecological risk for benthos in KODMDS.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 31 July 2017 Received in revised form 12 October 2017 Accepted 15 October 2017 Available online 17 October 2017

Kaohsiung Ocean Dredged Material Disposal Site (KODMDS) that located in the southwest offshore of Taiwan, has been annually disposed about 500,000 ton dredged sediments of Kaohsiung Harbor from 2003 to 2012. Five sediment cores collected from KODMDS and three from nearby reference sites were analyzed to evaluate their sedimentation rates, vertical profiles of heavy metal, and heavy metal pollution indices to assess the impact of dumping harbor dredged sediments into the ocean on the sediment quality in KODMDS. The sedimentation rate of 0.24 cm/y was estimated by the 210Pb method, which means that the effected depth of the top layer of a core of D1 was affected in the period of dumping dredging sediments. The vertical distribution of heavy metals in the sediment cores from KODMDS showed the concentrations of most heavy metals were slightly elevated in the top layers of the sediment cores, which may be affected by the dumping of harbor dredged sediments. According to the analyzed results of the heavy metal pollution indices, the level of heavy metal pollution, the potential eco-toxicity and the potential ecological risk of the sediments in KODMDS exhibited only a slight increase, which indicated that the increase in concentration of heavy metals may potentially pose the insignificant impact on benthos inhabiting the disposal site. © 2017 Published by Elsevier Ltd.

Handling Editor: J. de Boer Keywords: Heavy metal Kaohsiung Ocean Dredged Material Disposal Site (KODMDS) Sediment core

1. Introduction Harbor dredged sediments dumping to the given ocean disposal site may cause the adverse effects such as erosion and precipitation, deterioration of water quality (e.g. increase in turbidity and nutrients), and change in habitat, food chains, and fisheries on the marine environment and ecological system (Johnson and Frid, 1995; Zimmerman et al., 2003; Simonini et al., 2005; Lee et al., 2010). Fang et al. (2013) indicated that the major factors influencing the changes to benthos in the disposed site included particle

* Corresponding author. Department of Marine Environmental Engineering, National Kaohsiung Marine University, Kaohsiung, 81157, Taiwan. E-mail address: [email protected] (C.-D. Dong). https://doi.org/10.1016/j.chemosphere.2017.10.091 0045-6535/© 2017 Published by Elsevier Ltd.

size, water content of sediments, and water mass, while heavy metal concentration in sediments had no significant effect on the change. Stronkhorst et al. (2003) assessed the impact on the marine disposed site caused by dumping the moderately contaminated dredged sediments and indicated marine benthic organisms at and around the dumping sites have been adversely affected by physical disturbance such as burial and smothering. Furthermore, the toxic chemicals (e.g. heavy metals) in the dredged sediments may transport and accumulate in the marine environment (De Witte et al., 2016; Ju et al., 2016; Chen et al., 2017a). However, the previous studies demonstrated that the effect of ocean dumping was limited, in which may be caused by the below reasons: (i) the characteristics of dumping materials were similar with that of the sediments in the ocean disposal site; (ii) the sediments were dumped gradually and homogeneously over

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relatively large areas; (iii) using the way of dumping sediment was planning to disperse the dredged sediments on the seabed, enabling the benthic organisms to have the opportunity to migrate; (iv) the ocean disposal site was a high-energy environment that has the higher energy to carry and transport large particles, and the organisms that live in this area have adapted to be in the dynamic sediment environment for long time; (v) the level of the pollutant was low in the dredged materials (Smith and Rule, 2001; Simonini et al., 2005). Kaohsiung Ocean Dredged Material Disposal Site (KODMDS) was set up by Taiwan Environmental Protection Agency (TWEPA) in 2003, which was the given disposal place to dump the materials that dredged from the Kaohsiung Harbor for waterway maintenance. KODMDS was located in the southwest offshore of Taiwan with the distance of 12e15 miles from the shore, using longitude of E120 03.590 and latitude of N22 27.570 as the center and was the square area with the side length of 6 km (6  6 ¼ 36 km2). The water depth is range of 500e700 m in this area, and the change of water depth mainly shows the descending direction of the northeast to the southwest (Fig. 1). In the water depth of 200 m above, the directions and the average speed of ocean currents in the disposed site are mainly southeast and northwest and 10e40 cm/s, respectively; in the water depth of 200 m below, the directions of ocean currents is not constant and the speed of that is relatively low and decreasing with the increasing water depth (~0 cm/s). KODMDS has been annually disposed about 500,000 ton sediments that dredged from the channel of Kaohsiung Harbor, and it was of total about 4.69 million ton dredged sediments disposed to which area in 2003e2012 (TIPC, 2013). Kaohsiung Harbor endures the wastewater produced by neighboring industries, urban, and agriculture, as well as four main polluted rivers (Love River, Canon River, Jen-Gen River, and Salt River) import into this area, resulting harbor sediments contains high level of organic matter and heavy metals and belongs the moderate pollution level (Chen et al., 2013, 2016; Dong et al., 2014, 2015).

Dumping the contaminated sediments into the uncontaminated ocean disposal site was expected to have a certain degree of effect on that site. For preventing the concentrated dumping of dredged sediments and reducing the impact of dumping the dredged sediments on the ocean disposal site, the operation of ocean dumping has set the specific management strategy, including dividing the ocean disposal site into four sub-area to take turns dumping the dredged sediments, meaning each sub-area can only be dumped the dredged sediments of 15,000 ton and then turn to the other sub-area to dump. For the operation of dumping the dredged sediments, the dumping rate was kept to below 40 m3/min and the ship speed was controlled in the range of 1e4 knot. Each time to dump the dredged sediments was separated by the interval of more than 1 h (TIPC, 2005). This study sampled the sediment cores from KODMDS and reference sites and analyzed their physicochemical characteristics and heavy metals content to understand the vertical distribution of heavy metals and assess the impact of dumping the dredged sediments on the sediment quality in KODMDS, including the concentration of heavy metals, the degree of accumulation and the potential ecological risk. Furthermore, the effectiveness of operation management for dumping the dredged sediments can be determined based on the results of effect assessment of sediment quality. 2. Materials and methods 2.1. Sample collection Eight sampling sites were constructed to collect the sediments in KODMDS, including 5 disposed sites (D1eD5) inside KODMDS and 3 reference sites (R1eR3) outside (Fig. 1). Sites of D1eD4 were located in the center of four sub-areas inside KODMDS, D5 was in the center of whole KODMDS as well as R1eR3 were the north, east, and south reference sites located outside KODMDS. The sampling

Fig. 1. Map of the study area and sampling locations.

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method of sediment cores used the gravity-corer at the R/V Ocean Researcher III. The collected core samples were stored in the subcores (i.e., cylindrical plastic liners with a diameter of 7.3 cm) of the gravity-corer. After the core samples were transferred back to the laboratory, a meter ruler was used to measure the lengths of the sediment cores; subsequently, a stainless steel slicer was used to slice the cores at 2e5 cm intervals from top to bottom according to the core length. The sliced core samples were freeze-dried for 3 days and subsequently ground using a mortar and pestle. A mesh screen (0.063 mm) was used to sieve the ground samples to homogenize (Chen et al., 2016). The sediment samples after drying and homogenizing were contained in polyvinyl chloride bottles and then were prewashed with acid; sequentially, the pretreated samples were stored in a freezer at 20  C for further analysis.

index (Igeo) (Müller, 1979) and pollution load index (PLI) (Tomlinson et al., 1980) were applied to assessed the degrees of heavy metal contamination; the mean effect range median quotient (m-ERM-q) (Long et al., 2000) and potential ecological risk index (RI) (Hakanson, 1980) were employed to evaluate biological effects and potential ecological risks in the sediment cores. The detailed calculation process and definition of these indices see Chen et al. (2016). Moreover, all data were elaborated through principal component analysis (PCA), which is a particular type of factor analysis utilizing SPSS software (SPSS, version 12.0).

2.2. Grain size, total organic carbon and metal analysis

The vertical profile of 210Pb activity, both total and excess, in a disposed site (D1) and a reference site (R2) are shown in Fig. 2. The supported 210Pb activities, the near-uniform 210Pb activities at the core bottom were 3.32 ± 0.19 dpm/g for site D1 and 2.31 ± 0.45 dpm/g for site R2, respectively. The 210Pbex activity in the upper parts of the sediment core shows a simple exponential decrease throughout the cores. The sedimentation rates for D1 and R1 cores were 0.24 (r2 ¼ 0.996, p ¼ 0.003) and 0.09 cm/y, respectively (Fig. 2). According to the result of sedimentation rate (0.24 cm/y) for D1, the depth of 2.4 cm sediments on the top core may have corresponded to the period of dumping dredged sediments (nearly 10 years) and its metal concentrations may be influenced by the dredged materials. In the view of metal concentrations in D1, the distribution of Cd content was not complete agreement with the assumption (high level in depth 2.4 cm) because it increased upwards from the depth of 15 cm (Fig. 4). However, Cd in surface sediment indeed increased relatively sharply, which may be caused by dumping activity. But the increase in Cd content from the deeper depth may be produced by the local hydrologic and geochemical conditions or the body of plankton which contained Cd as a nutrient (Lin et al., 2002). Generally, the 210Pb method was not only applied to determine sedimentation rates in the top layer of sediment cores but also to discriminate the chronology construction of the sediments. Due to the dumping activity may make the disturbing in surface sediments, the present study only used 210Pb method for understanding the relative differences of sedimentation rates between the

The sediment samples were analyzed for grain size, and total organic carbon (TOC); the analysis procedure was listed in the previous study (Chen et al., 2012b). The method used to analyze the sediment core contents for heavy metals (Pb, Cd, Cr, Cu, Zn, Ni, Mn, and Fe) was as follows: 2.000 g of dry sediment sample was combined with mixed ultrapure acid (HNO3:HCl:HF ¼ 5:2:5, V/V/V) and then digested by a microwave digester (MARS 5, CEM, USA). The digested sample was filtered using a filter paper with a 0.45 mm pore size and the sample was quantitatively diluted to 15 mL by using ultrapure water. The concentrations of heavy metals were then measured using flame atomic absorption spectrophotometry (FAAS) (Hitachi Z-6100, Japan). The analysis of each batch was accompanied with an analysis of marine sediments reference materials for trace metals (PACS-2), blank samples, check samples, and duplicate samples. The difference between the sample analysis results and verified heavy metals values measured through the PACS-2 was smaller than 10%; the measured values of each heavy metal in the blank samples were lower than the detection limit values; the recoveries of the check samples ranged from 94.6% to 107.1%; and the percentage difference of all the duplicate samples ranged from 2.0% to 12.3%. The detection limit values (mg/kg dry weight) of metals were as follows: Cd (0.01), Cr (0.1), Cu (0.5), Pb (0.1), and Zn (0.8). 2.3. Radioactive dating

3. Results and discussion 3.1. Sedimentation rate

The 210Pb method was applied to determine sedimentation rates of the sediment cores in a disposed site (D1) and a reference site (R2). 210Po was measured by a-spectrometry to converse the activity of 210Pb (Institute of Oceanography, National Taiwan University), according to well-established methods constructed by Huh and Su (1999). The excess 210Pb (210Pbex) activity was derived by the total supported 210Pb subtracting the supported part which derived from the near-uniform 210Pb activities at the core bottom. The advection-diffusion model assumes that the sediment fluxes and the 210Pb activity reach the constant status given a long enough time and the excess 210Pb decreases with depth in the cores. Therefore, sedimentation rates can be estimated should be considered as maximum values under the assumption of ignoring the mixing process (Berger and Heath, 1968; Huh and Su, 1999). 2.4. Data analysis A statistical analysis method was employed for the data analysis (e.g., mean, standard deviation, and maximum and minimum concentrations). The Pearson correlation coefficients were used to test the relationship between sediment characteristics and metal concentrations. The enrichment factor (EF), geo-accumulation

Fig. 2. Vertical profiles of 210Pb activity (including both total and excess) in a disposed site (D1) and a reference site (R2).

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vertical distributions of particle composition in the sediment cores from the disposed sites (D1D5) display significant variation, while that from the reference sites exhibit a relatively constant (Fig. 3). The mean concentration and standard deviation of heavy metal contents in sediment cores collected from the disposed sites (D1D5) were as follows: Pb (15.6e18.5 mg/kg), Cd (0.03e0.05 mg/ kg), Cr (50.0e70.0 mg/kg), Cu (12.5e15.6 mg/kg), Zn (101.3e124.7 mg/kg), Ni (25.5e29.1 mg/kg), Mn (295e401 mg/kg), and Fe (2.3e3.5%). No significant difference was found in the mean metal concentration between all sediment cores (p < 0.05). Except for Cd, the concentration of heavy metals in the sediment cores of disposed sites was about two-fold higher than that of reference sites of R1 and R2 whereas was similar with that of reference site R3 (Table 1). Contrasting the source of dredged materials, the heavy metals concentration in sediments of Kaohsiung Harbor were 10e109 mg/ kg for Pb, 0.15e1.92 mg/kg for Cd, 11e523 mg/kg for Cr, 10e562 mg/ kg for Cu, and 60e1602 mg/kg for Zn, which was 5 times higher than the metal concentrations in sediments of the disposed site (Chen et al., 2012a). Comparing result found the metal concentrations in the sediment cores of disposed sites were closer with that in the sediment cores reference sites than that in dredged sediments of Kaohsiung Harbor. This phenomenon may be due to the fact that the dredged sediments had been diffused rapidly by strong ocean currents, or that the dredged sediments were constantly mixed with sea water, releasing the heavy metals inside while descending (the depth of KODMDS to between 500 and 700 m) (Tiller et al., 1989; Chen et al., 2017b). Galkus et al. (2012) indicated that the increased ocean current speed and active water circulation between the strait and the sea hinder the fine dredged particles to lodge in the bottom sediments in their study area. Especially the fine particles were recognized containing relatively higher metal concentrations. Barciela-Alonso et al. (2004) also reported that ocean currents play an important role in the distribution of metal concentration in sediments of Ria de Arousa, Spain. The vertical distribution of heavy metals in the sediment cores from KODMDS was shown in Fig. 4. The concentrations for most of the heavy metals were elevated in the top layers of the sediment cores in disposed sites, which may be affected by the dumping of dredged sediments. The concentration of Cd in sediment core of reference site R1 decreased with the depth from bottom to top, whereas the concentration of other metals slightly increased with the depth. For the reference site R2, the increase of metal concentrations was observed from the top portion to bottom layer in the sediment core. The concentration of all metals in R3 cores fluctuated with the depth, without showing obvious trends of increasing or decreasing. In the view of grain size distribution, for sediment

disposed site and reference site. However, the sedimentation rates were the preliminary evaluating based on the surficial sediment layers in this study, in which could be one of the estimates to prove the reference on the effect of dumping activity. Selvaraj et al. (2010) geochemically analyzed the sediment cores of dumping site in the southwestern Taiwan and indicated the sedimentation rates for the top and bottom portions of core were 0.19 and 0.74 cm/y, respectively. The difference between sedimentation rates demonstrated the dumping activities could influence the sedimentation on the top layer. This waste dumping site underwent 11 years of dumping activity, with 2 million tons of slag. Moreover, Selvaraj et al. (2010) observed the vertical distribution of metal concentration in the core and indicated that the thickness of 2.5 cm in sediments was resulting from slag dumping activity. The present study also obtained the consistent results with the report of Selvaraj et al. (2010). Huh et al. (2009) estimated sedimentation rates using 210Pb and 137 Cs as sediment chronometers in the continental margin offshore southwestern Taiwan. According to their data, the sedimentation rates of sampling sites close to D1 and R1 were 0.36 and 0.08 cm/y, respectively. The estimate of sedimentation rates in reference site was similar with that estimated in the present study, whereas that in the disposed site was higher than estimated in the present study. Noteworthy, Huh et al. (2009) collected these sediment cores in 2006, and KODMDS was only dumped for 3 years at that time. Therefore, the dumping activity was not the only reason for resulting in the relatively higher sedimentation rate in disposed site. The other reason may be KODMDS locating in the depocenter with high sedimentation rate flanking the Gaoping submarine canyon (Huh et al., 2009; Hsu et al., 2014). 3.2. Distribution of grain size, TOC and metal in sediment cores Table 1 lists the mean content of grain size (clay, silt, and sand), TOC and heavy metals concentration of the sediment cores collected from KODMDS (details as shown in Table S1). The mean percentage composition of clay, silt, and sand in sediment cores collected from the disposed sites (D1D5) were 10.4e11.7, 69.6e80.0, and 8.8e19.7%, respectively. Fine particle (<63 mm) was the dominant particles of sediments, and the variation in particle composition was mainly between sand and silt. The dominant particle composition in the reference site was also the fine particle, whereas the relatively higher composition of sand was found at sites R1 and R2, with 38.4 and 34.3%, respectively. The mean TOC of sediment cores collected from the disposed sites (D1D5) and reference sites (R1R3) were 0.93e1.08 and 0.65e0.94%, respectively. Fig. 3 shows the vertical distributions of particle composition and TOC in the sediment cores collected from KODMDS. The

Table 1 Characteristics and heavy metal concentrations in sediment cores from KODMDS. Itema

D1

D2

D3

D4

D5

R1

R2

R3

Sample size (n) Clay (<2 mm, %) Silt (2e63 mm, %) Sand (>63 mm, %) TOC (%) Pb (mg/kg) Cd (mg/kg) Cr (mg/kg) Cu (mg/kg) Zn (mg/kg) Ni (mg/kg) Mn (mg/kg) Fe (%)

21 11.7 ± 1.4b 78.3 ± 7.5 10.0 ± 8.7 0.93 ± 0.13 17.4 ± 0.8 0.04 ± 0.01 64.0 ± 5.9 13.6 ± 1.1 101.3 ± 8.3 26.9 ± 0.8 345 ± 19 2.8 ± 0.5

8 10.4 ± 1.8 80.0 ± 12.9 9.6 ± 14.6 1.06 ± 0.18 15.6 ± 0.7 0.05 ± 0.01 50.0 ± 4.8 12.5 ± 0.8 97.9 ± 9.1 25.8 ± 0.8 311 ± 40 2.3 ± 0.3

21 10.7 ± 3.2 69.6 ± 18.1 19.7 ± 21.2 0.94 ± 0.15 17.3 ± 0.8 0.05 ± 0.02 70.0 ± 11.9 13.8 ± 0.8 108.9 ± 9.0 27.1 ± 1.3 352 ± 26 3.5 ± 0.4

17 11.4 ± 2.3 74.2 ± 11.7 14.4 ± 13.7 1.06 ± 0.17 16.6 ± 0.8 0.04 ± 0.01 60.9 ± 6.8 12.8 ± 1.3 102.9 ± 8.3 25.5 ± 0.9 295 ± 24 2.5 ± 0.4

8 11.7 ± 1.5 79.6 ± 7.8 8.8 ± 9.2 1.08 ± 0.27 18.5 ± 1.8 0.03 ± 0.01 69.5 ± 5.1 15.6 ± 1.9 124.7 ± 7.3 29.1 ± 3.4 401 ± 53 3.0 ± 0.4

10 7.6 ± 0.8 54.1 ± 4.6 38.4 ± 5.2 0.80 ± 0.12 9.6 ± 0.8 0.07 ± 0.01 23.0 ± 1.1 5.5 ± 0.8 41.5 ± 2.9 16.7 ± 1.1 181 ± 17 1.3 ± 0.3

21 7.0 ± 1.6 58.7 ± 17.9 34.3 ± 19.5 0.65 ± 0.36 9.8 ± 2.6 0.01 ± 0.01 26.3 ± 9.0 7.7 ± 3.7 50.5 ± 12.8 19.8 ± 5.6 218 ± 69 1.6 ± 0.3

14 14.3 ± 1.7 83.8 ± 2.1 1.9 ± 3.1 0.94 ± 0.07 23.1 ± 1.9 0.04 ± 0.02 43.1 ± 4.1 17.8 ± 1.6 79.2 ± 3.6 29.4 ± 2.5 412 ± 38 2.6 ± 0.4

a b

All the values of item are basis on dry weight of sediment. Mean ± standard deviation.

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Fig. 3. Vertical distribution of particle size and TOC in sediment cores collected from KODMDS.

cores collected from disposed sites, there was no significant difference between in the top layer of core and in the remaining core layers. However, the vertical distribution of heavy metals in the sediment cores was in line with the estimated results of sedimentation rates, which the higher metal concentrations were found only on the top of core (Fig. 4). Selvaraj et al. (2010) also indicated that only a thin thickness (2.5 cm) of sediments was affected by a long-term and frequent dumping activity. Therefore, the vertical variation of the sand composition in the core collected from the disposed site may be caused by the local hydrological environmental conditions. Pearson correlation analysis was utilized to assess the relationships among metals concentration, grain size, and TOC in the sediment cores (Table 2). The significantly positive correlation (p < 0.01, r ¼ 0.62e0.99) was found among metals concentration, grain size, and TOC in reference sites except that Cd showed the quite low concentration in all sites. Results showed that metals concentration in the cores of reference sites was depended on grain size and TOC. Yet, sediment samples were sieved through a sieve to separate the sand-size particle (>63 mm) prior to analyzing heavy metals for preventing the influence of grain size in the present study. Lin et al. (2002) observed that the distribution patterns of heavy metal and organic carbon was similar with that of the portion of fine-grained sediments in the East China Sea continental shelf sediments. With a decrease in the percentage of fine grain, the sand-grained sediments increased accordingly. The sand-grained sediments were mainly constituted by quartz and/or biogenic carbonate and their existence diluted the heavy metal concentrations in sediments (Lin et al., 2002). Therefore, the distribution of metals concentration in sediments of reference sites may be dominated by the natural hydrologic and geochemical conditions. Notably, metal concentrations were not associated with grain

size and TOC in the sediment cores collected from the disposed sites. The present study extra and respectively analyzed the correlation between metal concentrations and grain size in the top layer and in the remaining layers of cores collected from the disposed site. No correlation was found in both analyses in the top layer and in the remaining layers (analyzed results not shown in the present study). In our expectation, the correlation between metal concentrations and grain size in the remaining layers of cores in the disposed sites should be significant as similar as which found in reference sites, owing to this part sediment cores were not associated with the dumping activity. According to the finding, in the area of disposed sites was influenced by other factors. In the continental margin offshore southwestern Taiwan, a pair of deposition lobes with high sedimentation rates flanks the Gaoping submarine canyon and may be led by sedimentation from turbidity flows overflowing the canyon (Huh et al., 2009; Hsu et al., 2014). KODMDS and reference site R3 exactly located on the one side of deposition lobes, therefore their vertical distribution of sediment texture was obviously distinguished with reference sites R1 and R2. In addition, Williams and Block (2015) reported that different hydrologic and geochemical conditions cause the varied MneFe distributions as well as the different metal contents, grain size, and TOC in marine sediments. Fig. 5 shows the MneFe distributions in the eight sediment cores collected from KODMDS. All data were significantly distinguished into two groups, one group including the reference sites R1 with all depth and R2 with a depth of 0e30 cm and the other group containing the rest of data. The clearly distinguishing in MneFe distributions between the two groups implied that the different hydrologic and geochemical conditions were the major drivers to make the difference in correlation among metals concentration of sediment cores. Results of MneFe distributions also indicated that the environmental

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Fig. 4. Vertical profiles of heavy metal concentration in sediment cores collected from KODMDS.

conditions of reference site R3 were similar with that of disposed sites. Overall, according to analyzed results in the present study, no matter for grain size, the distribution of metals concentration, and MneFe distributions, reference site R3 was similar with disposed sites comparing with R1 and R2. It was speculated that the relatively geological position was the main reason for this phenomenon due to that they all locate in the deposition lobes with high sedimentation rates flanking the Gaoping submarine canyon. The local ocean current and the sources of sedimentary particles may be

quite similar at R3 and disposed sites. Conversely, relatively farther distance between R1 and dumping area while R2 is located in the deeper water depth. Besides, the significantly positive correlation was observed among the contents of Pb, Cr, Cu, Zn, Ni, and Mn in the sediment cores of disposed sites, but no (p > 0.05) or low correlation (r ¼ 0.24e0.30) was found between Fe concentration and the other metal concentrations. It was speculated that part concentrations of heavy metals may be contributed from the same source such as

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Table 2 Pearson correlation matrix for metals concentration, grain size, and TOC in the sediment cores. Clay Reference site (n ¼ 45) Silt 0.86a Sand 0.90a TOC 0.63a Pb 0.97a Cd 0.38a Cr 0.89a Cu 0.94a Zn 0.90a Ni 0.88a Mn 0.93a Fe 0.86a Disposed site (n ¼ 75) Silt 0.89a Sand 0.92a TOC 0.03 Pb 0.22 Cd 0.11 Cr 0.06 Cu 0.13 Zn 0.00 Ni 0.11 Mn 0.13 Fe 0.00 a b

Silt

Sand

TOC

Pb

Cd

Cr

Cu

Zn

Ni

Mn

0.99a 0.83a 0.85a 0.22 0.94a 0.92a 0.91a 0.94a 0.90a 0.83a

0.81a 0.89a 0.25 0.95a 0.94a 0.93a 0.95a 0.93a 0.85a

0.60a 0.36b 0.75a 0.69a 0.69a 0.76a 0.68a 0.62a

0.31b 0.92a 0.97a 0.93a 0.90a 0.97a 0.88a

0.24 0.21 0.16 0.19 0.23 0.11

0.97a 0.95a 0.97a 0.96a 0.89a

0.98a 0.97a 0.98a 0.92a

0.96a 0.95a 0.91a

0.95a 0.90a

0.91a

0.99a 0.04 0.08 0.08 0.17 0.09 0.03 0.09 0.08 0.13

0.03 0.10 0.08 0.16 0.09 0.03 0.09 0.09 0.11

0.12 0.14 0.13 0.05 0.15 0.18 0.10 0.15

0.14 0.47a 0.88a 0.66a 0.57a 0.44a 0.10

0.09 0.26b 0.13 0.09 0.01 0.11

0.48a 0.66a 0.29b 0.49a 0.24b

0.68a 0.43a 0.39a 0.07

0.44a 0.53a 0.06

0.49a 0.28b

0.30a

Correlation is significant at the 0.01 level (2-tailed). Correlation is significant at the 0.05 level (2-tailed).

dumping dredged sediments on the top layer of the sediment cores in KODMDS was observed according to the studied results of sedimentation rate and the vertical distribution of metal concentrations. Hence, this study comparing and assessing the effect of dumping dredged materials on the sediments in the disposed site and their potential ecological risk by analyzing the metal concentration in the top layer of sediment cores.

Fig. 5. The scatter plot of MneFe distributions in the eight sediment cores collected from KODMDS.

dumping materials whereas Fe content may be also influenced by the ocean currents, diagenetic process, and environmental conditions. The association between Fe and other metal concentrations has been used to discriminate the contributions between natural levels and anthropogenic sources (Li et al., 2000). Presley et al. (1992) demonstrated that good correlation between Fe and metal concentration are expected for unpolluted sediments. As aforementioned, results of correlation analysis among heavy metals contents were obviously different in the disposed sites and reference sites, which may be caused by the effect of dumping dredged materials on the sediments in the disposed sites or by the differences of locally hydrologic and geochemical conditions. 3.3. Impact on sediment in the KODMDS This study used indices of EF, Igeo, PLI, m-ERM-q, and RI to assess the heavy metal pollution and potential ecological risk assessment in the sediment cores collected from KODMDS. The effect of

3.3.1. Enrichment factor (EF) The EF value is defined as the ratio of heavy metal concentrations in the sediment samples to that in the bottom slices of the studied cores (background concentrations), in which metal concentrations need to be normalized by dividing the respective concentration of reference elements (Chen et al., 2007; Selvaraj et al., 2010; Chen et al., 2016). Generally, the distribution of Fe concentration observed the peak concentration in the top layer and the decreased concentration in down-core was considered the occurrence of the diagenetic effect (Rajendran et al., 1992; Rowan et al., 2009; Tao et al., 2017). According to the vertical profile of Fe concentration in sediment cores shown in Fig. 4, it did not show that pattern, demonstrating that the influence of the diagenesis on the Fe concentration in cores may be insignificant. Concentrations of Al and Fe were used to be the normalized elements to prevent the metal variability associated with variations in different grain sizes in estimating EF for many studies because they are one of the major elements in the continental crust (Deely and Fergusson, 1994; Daskalakis and O'Connor, 1995; Lee et al., 2000; Abrahim and Parker, 2008). Sinex and Helz (1981) also mentioned that EF is not sensitive to the selection of normalized elements. Since Al content in the sediment cores was not measured in the present study, so Fe was used to be the candidate of reference metal. As shown in Table 3, the mean value of EF for 7 heavy metals in whole sediment cores were respectively in the range of 0.8e1.4 and 0.6e1.5 in the disposed sites and the reference sites, those belonged the classes between no enrichment to the minor. A recent study indicated that the mean EF of 1.0e4.2 in the ocean dumping site was relatively higher than that of 0.9e1.4 in the reference site in the similar region (Selvaraj et al., 2010). For the top layer of cores, EF

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Table 3 Pollution indices of EF, Igeo, PLI, m-ERM-q, and RI in the whole and top layer of sediment cores collected from KODMDS.a. Site

EF Pb

PLI

m-ERM-q

RI

0.6 0.9 0.6 0.6 0.6 0.4 1.2 0.6

1.0 1.0 0.9 1.0 1.0 1.0 0.7 1.1

0.18 0.17 0.19 0.17 0.20 0.09 0.11 0.18

19.3 19.0 20.5 18.8 18.8 16.7 9.1 21.6

0.5 0.8 0.6 0.3 0.9 0.1 1.4 0.4

1.4 1.1 0.9 1.3 0.9 0.9 0.6 1.1

0.22 0.17 0.19 0.20 0.20 0.10 0.08 0.17

32.5 24.0 19.5 26.3 17.3 9.4 7.1 22.0

Igeo Cd

(a) Whole core D1 1.3 0.9 D2 1.0 1.3 D3 1.1 0.8 D4 1.2 0.8 D5 1.0 0.9 R1 0.8 0.6 R2 0.8 1.1 R3 1.0 1.5 (b) Top layer of core D1 1.9 2.3 D2 0.8 1.6 D3 1.6 0.5 D4 1.7 1.9 D5 1.3 0.6 R1 0.8 0.1 R2 0.8 1.0 R3 1.3 1.8

Cr

Cu

Zn

Ni

Mn

Pb

Cd

Cr

Cu

Zn

Ni

Mn

1.3 0.9 1.3 1.3 1.2 0.8 0.8 1.1

1.4 1.0 1.1 1.2 1.1 1.0 0.7 1.0

1.3 1.0 1.1 1.2 1.1 0.8 0.9 1.0

1.3 1.0 1.1 1.2 1.0 0.8 0.8 1.1

1.2 0.8 1.1 1.1 1.0 0.9 0.8 1.0

0.5 0.6 0.6 0.5 0.6 0.5 1.1 0.6

1.1 0.3 1.2 1.0 0.9 1.0 0.7 0.0

0.5 0.7 0.4 0.4 0.4 0.6 1.1 0.5

0.5 0.7 0.6 0.5 0.6 0.3 1.3 0.6

0.5 0.6 0.7 0.5 0.6 0.5 1.0 0.6

0.6 0.6 0.7 0.5 0.6 0.5 1.1 0.5

2.0 0.9 1.3 2.1 1.6 0.7 0.8 1.3

2.8 0.8 1.8 2.1 1.4 0.9 0.6 1.2

2.0 0.8 1.4 2.0 1.3 0.7 1.0 1.0

1.4 0.8 1.3 1.6 1.2 0.7 0.8 1.2

1.5 0.7 1.3 1.8 1.0 0.9 0.8 1.4

0.2 0.6 0.3 0.3 0.6 0.3 1.4 0.5

0.2 0.4 1.9 0.2 1.6 3.2 1.1 0.0

0.1 0.5 0.6 0.1 0.3 0.5 1.4 0.4

0.5 0.6 0.1 0.0 0.5 0.1 1.8 0.6

0.0 0.7 0.6 0.1 0.5 0.4 1.2 0.8

0.5 0.6 0.6 0.5 0.7 0.4 1.5 0.6

a Classification of pollution indices: EF  1: no enrichment; 1e3: minor; 3e5: moderate; 5e10: moderately severe; 10e25: severe; 25e50: very severe; and 50: extremely severe (Birth, 2003). Igeo <0: none; 0e1: none to medium; 1e2: moderate; 2e3: moderately strong; 3e4: strong; 4e5: strong to very strong; and >5: very strong (Müller, 1979). m-ERM-q <0.1: 12% probability of toxicity; 0.11e0.5: 30% probability of toxicity; 0.51e1.5: 46% probability of toxicity; and >1.5: 74% probability of toxicity (Long et al., 1998). RI < 50: low ecological risk; 50e100: moderate ecological risk: 100e200: considerable ecological risk: and 200: very high ecological risk (Li et al., 2012).

values of 0.5e2.8 for the disposed sites (D1eD5) were also higher than those of 0.1e1.8 for the reference sites (R1eR3). In general, although the value of EF in the top layer of cores in the disposed sites was slightly higher than that in the reference sites, both of them were ranked in the classes between no enrichment to the minor. Analyzed results of EF indicated that the heavy metal concentrations of the surface sediments in the disposed site may increase slightly after 10 years of dumping dredged sediments, but the increase of metal contents did not bring the increasing rank of enrichment class. 3.3.2. Geo-accumulation index (Igeo) The Igeo value was calculated by the ratio of log2 heavy metal concentration in sediments to 1.5 times background metal concentration. Table 3 indicates that the mean Igeo values of 7 heavy metals collected from the disposed sites (D1D5) and reference sites (R1R3) are less than 0, meaning they belong to the uncontaminated class. For the top layer of cores, the Igeo value of reference sites (R1eR3) was in the range of 1.8e0.0 which belonging uncontaminated class. The Igeo value of disposed sites fell within the range of 1.9e0.5, in which Cd, Cu, and Zn for site D1 (Igeo ¼ 0.2, 0.5, and 0.0), Cd for site D2 (Igeo ¼ 0.4), and Cu for site D4 (Igeo ¼ 0.0) were the classes of none to medium, whereas the other ones belonged to the uncontaminated class. Results of Igeo implied that parts of heavy metals (Cd, Cu, and Zn) were accumulated in the top layer of disposed site due to dumping the dredged sediments that gained from Kaohsiung Harbor. 3.3.3. Pollution load index (PLI) PLI is the n-th root of the product of contamination factors in sediments (n for the species no. of studied metals). The value of contamination factor is the ratio of heavy metal concentration in sediments to the background metal concentration. The pollution load index (PLI) is a comprehensive assessment of the heavy metals pollution level. A PLI that is  1 or <1 indicates that sediment is contaminated or uncontaminated by heavy metals, respectively (Tomlinson et al., 1980). Table 3 lists the mean PLI values of sediment cores in KODMDS. The mean PLI values for whole sediment cores in the disposed sites (D1D5) and reference sites (R1R3) were all ranged of 0.7e1.1, and only site R3 was found to be the class

of contaminated. The PLI values were 0.9e1.4 for the top layer of sediment cores in the disposed sites, in which PLIs for sites D1, D2, and D4 were above than 1 for (PLI ¼ 1.4, 1.1, and 1.3), belonging to the contaminated level. The PLI value of the top layer of the core in the reference site R3 was 1.1, which was equal to the mean PLI of the whole sediment core. The reference sites R1 and R2 were evaluated as the uncontaminated level, with PLI values of 0.9 and 0.6, respectively. It can be seen that dumping the dredged sediments in KODMDS could result in some degree of heavy metal pollution. 3.3.4. Mean effect range median quotient (m-ERM-q) The m-ERM-q is the mean of ERM to sediment-based heavy metal concentration quotient and can be applied to assess the potential biological effects of the composite heavy metals in sediments (Gao and Chen, 2012; Chen et al., 2013). Table 3 shows the mERM-q distributions (calculated using Pb, Cd, Cr, Cu, Zn, and Ni) in sediment cores collected from the 8 sampling sites. The mean mERM-q values for the disposed sites (D1D5) were ranged of 0.17e0.20. Accordingly, the sum of the six heavy metals may achieve a 30% probability of toxicity (Long et al., 1998). The mean mERM-q values for the reference sites (R1R3) were ranged of 0.09e0.18. The probability of toxicity for sediments of sites R2 and R3 were ranked at the 30% probability of toxicity, while the sites R1 was ranked at the 12% probability of toxicity. For the analysis of mERM-q in the top layer of the sediment core, values of 0.17e0.22 were estimated in the disposed site, representing a 30% probability of toxicity was reached on the given site. The m-ERM-q values in the top layer of core were 0.10, 0.08, and 0.17 in reference sites R1, R2, and R3, respectively, in which sites R1 and R2 were the low probability of toxicity caused by heavy metals, while site R3 was the medium-low probability of toxicity (Table 3). Based on the hydrologic and geochemical conditions in the disposed sites similar with that in the reference site R3 (Fig. 4), the m-ERM-q values in the disposed site should be compared with that in the reference site. Therefore, although dumping the dredged sediments of Kaohsiung Harbor may slightly increase the m-ERM-q values in sediments, no significantly increased probability of toxicity was found. 3.3.5. Potential ecological risk index (RI) The estimate of RI is calculated as the sum of potential ecological

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risk coefficients for all studied metals, in which the potential ecological risk coefficient for single heavy metal is by the biological toxicity factor multiplying the ratio of heavy metal concentration in sediments to the background metal concentration. The RI value was also used to comprehensively assess the ecological risks caused by heavy metals (Hakanson, 1980). Classic PERI method considers eight pollutants, including PCBs, Hg, Cd, As, Pb, Cu, Cr, and Zn. However, the present study analyzed and considered 6 heavy metals of Pb, Cd, Cr, Cu, Zn, and Ni in sediment cores. Due to the difference in contaminant types and quantity, the present study used the adjusted grading standard of heavy metals’ ecological risk indices (Li et al., 2012; Jiang et al., 2014). Making the rounding digit of S Ti (biological toxicity factor; Pb ¼ 5, Cd ¼ 30, Cr ¼ 2, Cu ¼ 5, Zn ¼ 1, and Ni ¼ 5) as the lowest level limit of RI, and the remaining level limits followed by doubles (Li et al., 2012; Jiang et al., 2014). The adjusted grading standards of the potential risk of heavy metals in sediments were summarized in Table 3. Table 3 shows the distributions of the mean RI values of sediment cores for all studied sites calculating by six heavy metals (Pb, Cd, Cr, Cu, Zn, and Ni). The mean RI values for all disposed sites were lower than 50 (RI ¼ 18.8e20.5), indicating this area was exposed to a low ecological risk. The mean RI values for reference sites were between 9.1 and 21.6, hence the reference sites were also subjected to the low ecological risk. The RI values of the top layer of sediment cores in the disposed sites and reference sites were 19.5e32.5 and 7.1e22.0, respectively, both of them belonged to the low ecological risk. Although the increase of RI values was found in the top layer of sediment cores in the disposed sites D1, D2, and D4 (RI ¼ 32.5, 24.0, and 26.3), these sites were still classified to the low ecological risk, possibly meaning no increasing probability on ecological risk by dumping dredged materials. A variety of methods have been developed for the assessment of heavy metal risk. These assessing indices have been commonly used due to they will increase when sediment contamination increases (simple to quantify the pollution level) as well as they can be used providing a relative ranking of sampling sites. Especially, m-ERM-q and RI consider not only the total heavy metals concentration but also the ERM values and toxic-response factors are estimated by a large number of biological experimental data (Hakanson, 1980; Long et al., 1998). To take into account the natural background metal concentration, the estimates of EF, Igeo, and PLI are normalization by the metal concentrations in the bottom slices of the studied cores. These assessing indices have already been applied to study the contamination of heavy metals in sediments of the river, estuary, and marine environments (Long et al., 2000; Loska and Wiechuła, 2003; Zhang et al., 2009; Selvaraj et al., 2010; Gao and Chen, 2012; Chen et al., 2016). Although these indices could not accurately reflect on-site ecological risk caused by heavy metals, they may be useful for initial screening criteria and characterizing the suitability of the dredged material for disposal (Vidal and Bay, 2005). Furthermore, on-site ecotoxicological studies and benthic community analyses are needed for further investigation to clarify the linkage between the dumping dredged sediments in the ocean and the impact on which ecological system. 3.4. Principal components analysis PCA (the maximum varimax) was performed on all datasets obtained from sediment cores collected from the disposed sites and reference sites, the datasets comprising content of grain size (clay, silt, and sand), TOC, and heavy metals concentration. Based on result of PCA, two PCs (principal component) that could explain the total variation of 77.1% of the data in all sediment cores were obtained (Fig. 6(A)). PC1 could explain the variance of 39.8% of

563

Fig. 6. Rotated factor loadings (A) and scores (B) of PCs in the sediment cores from KODMDS.

sediment cores, in which the clay (0.90), silt (0.94), and sand (0.95) loadings were high, whereas Pb, Cu, and Ni and Mn had moderate loadings (loading >0.6). PC2 can explain 37.3% of the variance, with Cr (0.90), Zn (0.87), and Fe (0.78) exhibiting relatively high loadings and Pb, Cu, Ni, and Mn showing moderate loadings. Fig. 6(B) presents the scatter plot for the PC1 and PC2 of the top layer of core and remaining layers in the disposed sites and reference sites. The score of PC1 was depended on the distribution of grain size and the concentration of part heavy metals in sediments, whereas that of PC2 was mainly decided on the concentration of heavy metals. The difference of mean score of PC1 between top layer of core and the remaining layers for each sampling site was as follows from high to low: D3 (1.35) > R2 (0.92) > R1 (0.45) > D1 (0.37) > D4 (0.27) > D5 (0.23) > D2 (0.13) > R3 (0.07). The difference was mainly influenced by the distribution of grain size. As for PC2, the difference of mean score loading between the top layer of core and the remaining layers for each sampling site was as follows: D1 (1.83) > D4 (1.10) > D3 (0.81) > D2 (0.35) > D5 (0.21) > R1 (0.19) > R2 (0.06) > R3 (0.11), indicating the surface sediments were apparently affected by the dredged sediments and resulted that its metal concentration was higher than the mean metal concentration of the whole core. Nevertheless, according to the results of EF, Igeo, PLI, RI, and m-ERMq, an only slight increase in the heavy metal pollution, biological toxicity, and potential ecological risk were estimated, and the increase in these indices did not pose the significant impact on the ecological environments.

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4. Conclusions The analysis of physicochemical characteristics, sedimentation rate, and pollution indices was estimated in sediment cores of the disposed sites and reference sites in the present study, and their results showed that dumping the dredged materials from Kaohsiung Harbor could cause a slight increase of contamination in sediments of KODMDS. However, according to the determined indices, studied results also found that the slight increase of contamination in sediments did not impact on the sediment quality and the benthic ecosystem in the disposed site. Considering the difference in hydrological and geochemical conditions in the studied region, the sediment cores was used to correctly assess the effect that induced by man-made activity (i.e. dumping dredged materials) on the marine sediments, because the bottom of the core can be as the background to adjust the pollution estimates. About 500,000 ton dredged sediments of Kaohsiung Harbor were disposed to the KODMDS per year, and the dredged sediments with the moderated level of pollution, ecological toxicity, and risk would be transformed to the that with the low level of pollution and potential ecological risk via the diffusion of ocean currents and dilution effect. Hence, based on the results of this study, the ocean dumping may be a feasible way to solve a large number of dredged sediments from harbor under the effective management. In addition, the assessment conclusions of this study would be beneficial for the management and control of heavy metal pollution in sediments. Yet, on-site ecotoxicological studies and benthic community analyses are needed for further investigation to elucidate the linkage between the dumping dredged sediments in the ocean and the impact on which ecological system. Acknowledgements This work was supported by the Taiwan International Ports Corporation (TIPC) (100-V-222-E), Taiwan. The authors are sincerely thankful to Professors C.C. Su and Y.J. Yang of National Taiwan University for supporting radioactive dating and fabricating bathymetric chart, respectively,in this study. Appendix A. Supplementary data Supplementary data related to this article can be found at https://doi.org/10.1016/j.chemosphere.2017.10.091. References Abrahim, G.M.S., Parker, R.J., 2008. Assessment of heavy metal enrichment factors and the degree of contamination in marine sediments from Tamaki Estuary, Auckland, New Zealand. Environ. Monit. Assess. 136, 227e238. ns, P., Regueira-Miguens, M.E., BermejoBarciela-Alonso, M.C., Pazos-Capea Barrera, A., Bermejo-Barrera, P., 2004. Study of cadmium, lead and tin distribution in surface marine sediment samples from Ria de Arousa (NW of Spain). Anal. Chim. Acta 524, 115e120. Berger, W.H., Heath, G.R., 1968. Vertical mixing in pelagic sediments. J. Mar. Res. 26, 134e143. Birth, G., 2003. A scheme for assessing human impacts on coastal aquatic environments using sediments. In: Woodcoffe, C.D., Furness, R.A. (Eds.), Coastal GIS 2003. Wollongong University Papers in Center for Maritime Policy, Australia, 14. Chen, C.F., Chen, C.W., Dong, C.D., Kao, C.M., 2012a. Assessment of toxicity of polycyclic aromatic hydrocarbons in sediments of Kaohsiung Harbor. Taiwan. Sci. Total Environ. 463e464, 1174e1181. Chen, C.F., Dong, C.D., Chen, C.W., 2013. Evaluation of sediment toxicity in Kaohsiung harbor. Taiwan. Soil. Sediment. Contam. 22, 301e314. Chen, C.F., Ju, Y.R., Chen, C.W., Dong, C.D., 2016. Vertical profile, contamination assessment, and source apportionment of heavy metals in sediment cores of Kaohsiung Harbor, Taiwan. Chemosphere 165, 67e79. Chen, C.F., Chen, C.W., Chen, T.M., Ju, Y.R., Chang, Y.K., Dong, C.D., 2017a. Phthalate ester distributions and its potential-biodegradation microbes in the sediments of Kaohsiung Ocean Dredged Material Disposal Site, Taiwan. Int. Biodeterior. Biodegradation 124, 233e242.

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