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j.chemosphere.2017.10.110

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Accepted Manuscript
Bioavailability and toxicity of trace metals (Cd, Cr, Cu, Ni, and Zn) in sediment cores
from the Shima River, South China
Lei Gao, Zhuowei Wang, Shaoheng Li, Jianyao Chen
PII:
S0045-6535(17)31695-8
DOI:
10.1016/j.chemosphere.2017.10.110
Reference:
CHEM 20132
To appear in:
ECSN
Received Date: 11 June 2017
Revised Date:
17 October 2017
Accepted Date: 20 October 2017
Please cite this article as: Gao, L., Wang, Z., Li, S., Chen, J., Bioavailability and toxicity of trace metals
(Cd, Cr, Cu, Ni, and Zn) in sediment cores from the Shima River, South China, Chemosphere (2017),
doi: 10.1016/j.chemosphere.2017.10.110.
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Bioavailability and toxicity of trace metals (Cd, Cr, Cu, Ni, and Zn) in sediment
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cores from the Shima River, South China
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Lei Gao, Zhuowei Wang, Shaoheng Li, Jianyao Chen*
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School of Geography and Planning, Sun Yat-Sen University, Guangzhou, 510275, China
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Abstract: Five sediment cores (S1−S5) were collected from the Shima River to determine the
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bioavailability of trace metals (Cd, Cr, Cu, Ni, and Zn) using the modified European Community
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Bureau of Reference (BCR) procedure. The toxic effects of polluted sediment were assessed using
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the LUMISTox® bioassay with Vibrio fischeri and chemical models such as the toxicity unit (TU)
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of each metal, sum of all TUs (∑TU), and toxic risk index (TRI). The results showed that Cd, Ni,
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and Zn were present mainly in the acid-soluble and residual fractions, and the residual fraction of
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Cr accounted for the majority of the metal content (44%), while Cu was present mainly in the
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reducible and residual fractions. Cd had a mean enrichment factor (EF) of 15.1 and was
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considered to be severely enriched, while there was a minor enrichment of Cr and moderately
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severe enrichment of Zn, Cu, and Ni. From the LUMISTox® bioassay, an acute TU (TUa) value
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exceeding 0.4 was found at the upper and middle reach sites and was considered to represent
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slightly acute toxicity, whereas little acute toxicity was found at the lower reach site. The
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acid-soluble fraction of trace metals was the geochemical fraction mainly responsible for the acute
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toxicity of the sediment, and acid-soluble Zn and Ni were identified as important contributors to
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sediment toxicity.
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Keywords:
Riverine
sediment;
trace metal; geochemical fraction; toxic effects; Vibrio fischeri.
*
Corresponding
author. Tel.:+86
20 84115930.
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E-mail address: chenjyao@mail.sysu.edu.cn (J. Chen)
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1 Introduction
Riverine sediment is an important sink for trace metals in river basins because of the deposition
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of suspended or dissolved metals inputted by surface runoff and direct anthropogenic discharges
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(Nelson and Booth, 2002; Fu et al., 2014). However, riverine sediment can also be a source of
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trace metals for river water as a result of environmental changes (e.g., pH and redox) at the
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sediment-overlying water interface. For example, a decrease in redox at the interface between
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solid and liquid phases would facilitate reductive dissolution of Fe and Mn oxides, which lead to
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the release of trace metals that were bound to them (Mukwaturi and Lin, 2015). And low pH
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reduces the negative surface charges of sediment particles and Fe and Al oxides, promoting the
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solubility, mobility and bioavailability of metals co-precipitated with carbonates and sulfides (Du
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Laing et al., 2009). Generally, more than 90% of trace metals are bound to suspended solids and
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sediments, leading to their significant accumulation and enrichment in sediment in aquatic
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systems (Wei et al., 2016). Sediment pollution by trace metals is a challenging pollution issue due
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to their stability, non-degradation, persistence, bioaccumulation, and especially toxicity
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(Paramasivam et al., 2015; Gao et al., 2016).
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High concentrations of trace metals may cause sediment toxicity and have a huge impact on the
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survival and growth of benthic organisms (Roman et al., 2007). Therefore, chemical models such
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as the sum of all toxicity units (∑TU) and the probable effect level (PEL) from the sediment
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quality guidelines (SQGs) mainly based on the total concentration of trace metals are used to
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assess the sediment toxicity caused by trace metals (MacDonald, 1994; Pedersen et al., 1998;
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Zhang et al., 2016). However, Sauvé et al. (2000) suggested that the total concentration of trace
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metals cannot be used to assess short-term ecological risks, because it does not reflect the mobility,
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reactivity, or bioavailability of specific potentially toxic metals. The mobility and bioavailability
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of trace metals in sediment were given a priority for risk and toxicity assessment because of the
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bioaccumulation and biomagnification of trace metals up the food chain, severely threatening
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human health (Varol and Şen, 2012; Pejman et al., 2015). In fact, the mobility and bioavailability
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of the different geochemical fractions of trace metals in sediment can vary, thus the chemical
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partitioning of trace metals is typically and widely undertaken using the European Community
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Bureau of Reference (BCR) method, which is a sequential extraction procedure, to determine their
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environmental hazards and ecological risks (Zhang et al., 2017). According to the BCR method,
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the acid-soluble, reducible, and oxidizable fractions show a tendency of decreasing bioavailability,
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while the residual fraction is not bioavailable (Rodriguez et al., 2009; Nemati et al., 2011; Zhang
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et al., 2016).
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Ecotoxicity tests or bioassays can reveal the biological responses to the different stresses
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associated with trace metals (Zabetoglou et al., 2002; Roig et al., 2015). In a complex
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environmental matrix, such as sediment containing a diverse and complex range of contaminants,
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a toxicity assessment based on physicochemical parameters may yield anamorphic results due to
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the lack of both environmental information and consideration of the interactions among toxic
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substances. Nevertheless, bioassays can directly indicate the integrated effects (e.g., antagonistic,
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additive, and synergistic) of chemicals, overcoming the deficiencies in physicochemical
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assessments for establishing toxic risks. The luminescent bacteria Vibrio fischeri has been widely
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used as a sensitive tool for monitoring soil and sediment toxicities and is often selected as the first
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option in a battery of tests, considering speed and cost (Parvez et al., 2006; An et al., 2012;
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Serafim et al., 2013). Cooper et al. (2009) found that the interactive effects among Zn, Pb, and Cu
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on the bioluminescence of V. fischeri were synergistic. Thus, bioassays can provide a direct
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quantitative measure of the actual sediment toxicity.
The Shima River is a typical urbanized catchment and is located near an important water source
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area in the Pearl River Delta, South China. Since the 1990s, the Shima River water quality began
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to deteriorate, and the riverine sediments were severely polluted by Cd, Cr, Cu, Ni, and Zn, which
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are derived mainly from industrial effluents and agricultural runoff
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2017). The spatial distribution of trace metals shows significant differences as a result of vertical
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heterogeneities in sediment profiles (Gao et al., 2016). To some degree, riverine core sediment can
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reflect the sedimentary dynamic process of former conditions, and the vertical distribution of
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geochemical fraction of trace metal provides intuitive perspectives to reveal variation trends of
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accumulation level and metal bioavailability, accurately assessing sediment pollution status caused
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by trace metals. However, little information is available regarding the geochemical fractions and
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toxic effects of trace metals in sediments from the Shima River catchment. Therefore, the aims of
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this study were to (1) quantitatively characterize the bioavailability and potential mobility of trace
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metals (Cd, Cr, Cu, Ni, and Zn) in sediment cores using the BCR sequential extraction procedure;
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(2) assess the sediment toxicity using a bioassay (luminescent bacteria) and chemical models, and
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establish the relationship between the observed and estimated toxic effects; and (3) preliminarily
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identify hazardous geochemical fractions and toxic elements.
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(Gao et al., 2015; Gao et al.,
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2 Materials and methods
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2.1 Study area
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The Shima River, with a length of 88 km and catchment area of 1249 km2, is located in
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Dongguan city, which is an important industrial and agricultural base in China. The study area,
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covered by unconsolidated quaternary sediments with lateritic red soil, has a subtropical monsoon
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climate and experiences an annual mean precipitation of 1954 mm and a mean temperature of
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22.9°C. The Shima River is the largest tributary of the Dongjiang River in Dongguan city and
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originates from the Guanlan River in the Baoan district of Shenzhen city (Fig. 1). The river water
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flows from south to north and subsequently discharges into the Dongjiang River, which is an
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important water source for urban areas. However, during the last 3 decades, residential population
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increased by a factor of approximate 5, simultaneously, five pillar industries including food
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processing, wood processing, paper-making, and metallic equipment and machinery
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manufacturing industry experienced a rapid development in Dongguan city, resulting in an
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average wastewater discharge volume of 13.6 × 108 m3 from 2005 to 2013 (Dongguan Water
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Authority, 2005-2013). Furthermore, to promote local agricultural productions, the cumulative
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application of fertilizers and agricultural chemicals were 5860 and 159 kg·ha-1 during the period of
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1982−2013 (Dongguan Bureau of Statistics, 1978-2013), respectively, directly leading to Cd
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accumulation in farmland along the river bank (Gao et al., 2015). The Shima River served as a
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receiving water body for industrial effluent, domestic sewage and agricultural runoff. Currently.
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Since 2003, the river water is obstructed by a rubber dam (height, 3.25 m and length, 92 m) to
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prevent contamination of the inlet of the water supply pumping station during the dry season,
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whereas the river drains into the Dongjiang River during heavy rainfall. Therefore, the Shima
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River catchment is of strategic importance for the protection of the local water source as a result
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of the relatively short distance between the inlet and outlet of the drainage from the Shima River.
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Fig. 1 Location of the study area and sampling sites.
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2.2 Collection of sediment cores
Five sediment cores were collected at the center of river channel from the upper (S1 and S2),
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middle (S3 and S4), and lower (S5) reaches of the Shima River (Fig. 1) in April 2015. The cores
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were taken using a gravity sampler equipped with a polycarbonate tube, with an inner diameter of
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6.5 cm and a length of 80 cm. After sampling, the overlying water was siphoned off using a tubule
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and syringe, and a stainless steel tapping pin with a degree scale was used to push the sediment
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out of the pipe from the bottom. Sediment cores S1, S2, S3, and S5, with a total length of 15 cm,
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were sectioned at an interval of 1 cm. For core S4, with a total length of 30 cm, an interval of 2 cm
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was applied. All sub-samples were separately sealed in polyethylene bags and stored at −20°C in
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the dark after transport to Guangdong Provincial Key Laboratory of Environmental Pollution
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Control and Remediation Technology, Sun Yat-Sen University.
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2.3 Physicochemical analysis of sediment samples
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After freeze-drying, sediment samples were invariably stored in refrigeration (−20°C) under
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dark conditions unless physicochemical pretreatment or analysis. One portion of the sample was
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ground with an agate mortar pestle and homogenized by straining through a 1-mm nylon sieve.
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The sediment pH was measured in a 1:2.5 (w/v) ratio of dry sample-to-deionized water with a
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water quality analyzer (LAQUAtwin; Horiba, Shanghai, China). The rest part of sample was
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ground using a mortar and pestle until all particles could pass through a 100-mesh (corresponding
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to 0.149-mm) nylon sieve (Yang et al., 2013; Zhang et al., 2017), to determine the sediment
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organic carbon (OC) content, total concentration and geochemical fraction of trace metals. The
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OC content was determined by dichromate oxidation (Nelson and Sommers, 1982).
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Approximately 50 mg sample was digested in a HNO3–HF mixture in Teflon tubes, and the trace
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metal (Cd, Cr, Cu, Ni, Zn, and Sc) concentrations were analyzed by inductively coupled plasma
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(ICP) mass spectrometry (X-2, Thermo Fisher Scientific, Waltham, MA, USA). Quality assurance
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and quality control were assessed in each batch of samples using a method blank and the stream
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sediment standard reference material GBW07312 (GSD12) from the Institute of Geophysical and
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Geochemical Exploration, Chinese Academy of Geological Science (Han et al., 2015; Zhang et al.,
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2016). The concentrations of recovered trace metals in the standard material ranged from 94.3 to
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101% (Table S1). Furthermore, 10% of all sediment samples were randomly selected for duplicate
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analysis, and the standard deviations of the duplicate samples were <10%.
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The geochemical fractions of trace metals in sediment samples were analyzed according to the
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modified BCR method, which involves a three-step sequential extraction procedure (Rauret et al.,
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1999). Full details of the modified BCR method are published elsewhere (Liu et al., 2010; Li et al.,
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2015). Briefly, four fractions of the trace metals (Cd, Cr, Cu, Ni, and Zn) were extracted from 0.5
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g sediment sample as follows: (F1) acid-soluble fraction (20 mL 0.11 mol·L-1 CH3COOH, shaken
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for 16 h), (F2) reducible fraction (20 mL 0.5 mol·L-1 NH2OH·HCl at pH 1.5, shaken for 16 h), (F3)
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oxidizable fraction (10 mL 0.88 mol·L-1 H2O2 + 25 mL 1.0 mol·L-1 CH3COONH4 at pH 2, shaken
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for 2 h + 16 h), and (F4) residual fraction, which was determined as the first three fractions (F1 +
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F2 + F3) obtained from the BCR extraction subtracted from the total concentration of trace metal
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(Kumar et al., 2013). The first two steps and third step of the extraction process were conducted in
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a water bath oscillator at constant temperatures of 25 ± 1°C and 90 ± 5°C, respectively. After each
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step of the extraction, separation of the liquid and solid phases was achieved by centrifugation of
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the mixtures at 4000 g for 20 min, and the supernatant was then filtered through a polypropylene
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membrane with a diameter of 0.45 µm for the measurement of Cd, Cr, Cu, Ni, and Zn
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concentrations by ICP-atomic emission spectrometry (IRIS, Pleasant Prairie, WI, USA). The
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corresponding detection limits were 2, 20, 4, 10, and 10 µg·L-1, respectively. Quality assurance
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and quality control were assessed using the lake sediment standard reference material (BCR-701)
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from European Community Bureau of Reference, and the recoveries of F1, F2, F3, and total
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concentration from the standard material ranged from 101 to 117%, 81.4 to 101%, 97.0 to 117%,
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and 88.4 to 109%, respectively (Table S2).
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2.4 Enrichment factor (EF)
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The EF, which normalizes the concentration of a trace metal to that of a conservative element,
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has been used to describe anthropogenic enrichment quantitatively (Audry et al., 2004; Bing et al.,
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2016). In the present study, Scandium (Sc) was selected as a reference element due to its wide
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distribution in crustal rocks and scarcity in various pollution sources (Steinmann and Shotyk, 1997;
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Gao et al., 2017). The EF is calculated using the following equation (Bing et al., 2016):
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EF = (
()
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)
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where (Me/Sc)s is the ratio of the concentration of a given metal (Me) to that of Sc in a sediment
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sample, and (Me/Sc)B is the soil background ratio between that metal and Sc in Guangdong province
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(China National Environmental Monitoring Centre, 1990). Enrichment levels were classified into five
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categories based on EF values: (1) EF < 1, no enrichment; (2) EF 1–3, minor enrichment; (3) EF 3–5,
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moderate enrichment; (4) EF 5–10, moderately severe enrichment; (5) EF 10–25, severe enrichment;
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(6) EF 25–50, very severe enrichment and (7) EF > 50, extremely severe enrichment (Chen et al.,
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2007).
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2.5 The TU and toxic risk index (TRI)
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SQGs were applied for qualitative assessment of the potential ecological risks caused by Cd, Cr,
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Cu, Ni, and Zn in the sediment. The SQGs have two thresholds: the threshold effect level (TEL),
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below which adverse effects on aquatic ecosystems are not expected to occur, and the probable
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effect level (PEL), above which adverse effects are likely to be observed (MacDonald, 1994; Esen
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et al., 2010). The ∑TU was used for quantitative determination of the toxic effects. The ∑TU was
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defined by Pedersen et al. (1998), who suggested that the potential acute toxicity of heavy metals
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in sediment could be represented by the sum of the ratio of the measured metal concentration to
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the PEL. The sediment ∑TU was calculated using the following equation:
∑TU = M
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However, the sediment ∑TU might underestimate the potential toxicity, because it only
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considers the PEL effects; therefore, the TRI was used to assess the integrated toxic risk based on
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both the TEL and PEL effects of trace metals (Zhang et al., 2016). The TRI of the sediment was
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calculated using the following equation:
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TRI = ∑ TRI = TE
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( /"#$ )% &( /'#$ )%
,
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(
where Csi is the concentration of metal i (mg·kg-1) in the sediment sample, and CiTEL and CiPEL
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are the TEL and PEL of metal i (mg·kg-1), respectively. Toxic risks were classified into five
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categories based on the TRI value: (1) TRI <5, no toxic risk; (2) TRI 5–10, low toxic risk; (3) TRI 10–
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15, moderate toxic risk; (4) TRI 15–20, considerable toxic risk; and (5) TRI >20, very high toxic risk.
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2.6 Bioavailable metal index (BMI)
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Sediment samples were analyzed to determine the geochemical partitioning of the trace metals Cd,
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Cr, Cu, Ni, and Zn according to the modified BCR method. F1, the most bioavailable fraction in the
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sediment, could be determined quantitatively. An integrated BMI was used to assess the total
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bioavailability of the investigated metals trapped in sediment. The BMI was calculated as follows
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(Rosado et al., 2015):
(
/
+
+
+
+
BMI = ( × (
× /
×∙∙∙× )
,-+
,-+
,-+ ,-+
where n is the number of trace metals investigated, and CiF1 and CiB-F1 are the concentrations of
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metal i in the F1 of the sediment and background samples, respectively.
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2.7 Toxicity assay using V. fischeri
A solid-phase toxicity assay was performed. Approximately 1.5 g freeze-dried sediment was
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resuspended in 15 mL 2% NaCl prepared with 0.1 mol·L-1 Na3PO4 at pH 7.0 ± 0.1 for 30 min in an
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oscillator at a constant temperature of 25 ± 1°C. A previous study suggested that phosphate buffer
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can prevent significant variations in pH and can be used to regulate bacterial metabolism
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(Salizzato et al., 1998). The supernatant was filtered through a 0.45 µm polypropylene membrane
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for the acute toxicity assay using the luminescent bacteria.
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LUMISTox® 300 (HACH LANGE GmbH, Duesseldorf, German) with V. fischeri was employed
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to assess the acute toxicity of supernatant prepared from the resuspended sediment using a method
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published by ISO (1999). The method consists of two steps. First, freeze-dried luminescent
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bacteria are reactivated after injecting a reconstitution solution (1 mL) and then transferred from
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−18℃ to 4℃ in a refrigeration unit prior to use. The luminescent bacterial suspension was diluted
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with 2% NaCl to meet the lowest emission intensity of 400 luminescent units and, after
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approximately 15 min, was then used in the assays. Second, the acute toxicity of the supernatant
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was assessed with the luminescent bacterial solution (0.5 mL) in a 1:1 (v/v) ratio, with each sample
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(0.5 mL) placed into a cuvette at 15℃. Toxicity assays of each sediment sample were performed in
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triplicate. Each supernatant sample was diluted with an equal volume of 2% NaCl to determine the
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dose–response relationship for calculation of the effective concentration causing a 50% reduction
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in bioluminescence intensity (EC50), if the luminescence inhibition rate (LIR, %) exceeded 60%.
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The LIR of V. fischeri and the relative acute TU (TUa) were obtained after a 30 min incubation
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according to the following equations (Libralato et al., 2010; Usero et al., 2016):
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LIR = (Ic/Is – 1) × 100%
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TUa = LIR × 2
(LIR ≤ 50%)
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TUa = 1/EC50
(LIR > 50%),
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where Ic and Is are the bioluminescence intensities of the control and supernatant sample,
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respectively. Toxicity classification was determined according to the TUa: TUa < 0.4, no acute
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toxicity; 0.4 < TUa < 1, slight toxicity; 1 < TUa <10, acute toxicity; 10 < TUa <100, high toxicity;
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and TUa > 100, very high toxicity (Libralato et al., 2010).
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2.8 Statistical analysis
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One-way analysis of variance (ANOVA) with Duncan's test was applied to determine the
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significant differences among the total concentration of trace metal and TUa values of the
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sediment cores, and p <0.05 (two-tailed) was considered to indicate a significant difference.
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Determination coefficient (R2) was derived from Pearson correlation analysis. Principle
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component analysis (PCA) was applied to identify potentially toxic elements in riverine sediment
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cores. The validity of PCA was checked using the Kaiser-Meyer-Olkin (KMO) statistic and
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Bartlett's test; a KMO value of p < 0.5 and a significant Bartlett's test value were prerequisites. All
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statistical analyses were carried out using SPSS ver. 16.0 for Windows. Linear fitting including 95 %
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confidence band and prediction band were completed by SigmaPlot ver. 10.0.
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3 Results and discussion
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3.1 Sediment property
Sediment property such as pH value and OC, etc. plays an important role in affecting the
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mobility and bioavailability of trace metals in riverine sediment. As shown in Fig. S1, the vertical
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distributions of pH value and OC content showed obvious differences among sampling sites. On
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the whole, sediment environment was slightly acid with a mean pH value of 6.24 (range, 4.90–
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7.30). The average OC content was 36.1 g∙kg-1 for S1, 13.6 g∙kg-1 for S2, 33.6 g∙kg-1 for S3, 39.6
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g∙kg-1 for S4, and 3.91 g∙kg-1 for S5, respectively. It was reported that the discharge of domestic
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wastewater and industrial effluent contributed significantly to sediment OC in the study area (Gao et al.,
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2017). And OC content exceeding 10 g∙kg was considered to reach to the lowest effect level (LEL),
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and may impose adverse effects on the sediment-dwelling organisms (Ontario Ministry of Environment
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and Energy, 1993).
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3.2 Total concentrations and geochemical fractions of trace metals
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The total concentrations of trace metals in sediment cores from the Shima River ranged from
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0.204 to 1.91 mg·kg-1 for Cd, 13.8 to 469 mg·kg-1 for Cr, 12.0 to 630 mg·kg-1 for Cu, 14.1 to 253
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mg·kg-1 for Ni, and 57.6 to 1700 mg·kg-1 for Zn (Table 1 and Fig. 2). Higher concentrations of
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investigated trace metals with exception of Sc were found at sites in the middle reaches (S3 and
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S4) than at sites in the upper (S1 and S2) and lower reaches (S5) (p <0.05), and the Cr, Ni, Cu, Zn,
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and Cd concentrations in the top sediment layer (0–10 cm) in cores S1−S4 exceeded the TEL or
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PEL values, resulting in large ecological risks. In core S5, the total concentrations of trace metals
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were below or close to the TEL values (Table 1), indicating only a limited potential for adverse
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ecological effects. In general, the total concentrations of trace metals decreased with depth in the
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sediment cores, except for core S5, in which metal concentrations remained relatively constant.
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Table 1 Statistical summary of heavy metal concentrations in different sediment cores from Shima River (mg·kg-1)
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Geochemical partitioning is a more useful measure than the total concentration of trace metals
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in sediment in terms of understanding their bioavailability, potential sources, and environmental
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behaviors. The geochemical fractions of trace metals based on a modified BCR sequential
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extraction procedure are shown in Fig. 2. The different geochemical fractions of Cd, Ni, and Zn in
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the sediments had a similar distribution. F1 accounted for 37.5%, 35.0%, and 51.6% of the total
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concentrations of Cd, Ni, and Zn, respectively, while F4 accounted for 40.3%, 32.4%, and 23.3%
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of the total concentrations of these metals, respectively. The mean proportions of the various Cu
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fractions followed the decreasing order of F2 (35.0%) > F4 (28.1%) > F1 (21.0%) > F3 (15.8%),
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while Cr was unique, with large proportions of the total concentration present in F2 (mean:
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27.2%), F3 (26.7%), and F4 (44.0%). F1 is the most mobile and bioavailable fraction. The
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proportions of trace metals in this fraction followed a decreasing order of Zn (51.6%) > Cd
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(37.5%) > Ni (35.2%) > Cu (21.0%) > Cr (2.09%), indicating that Cd, Ni, and Zn were very
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mobile and had the potential to be assimilated by benthic organisms and to impose environmental
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hazards on aquatic ecological systems (Zhuang et al., 2016). However, the potential bioavailability
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of sediment metals derived from the first step of the BCR sequential extraction was likely
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overestimated. Previous studies indicated that bioavailable fraction of sediment metals extracted
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by proteinase K was better linearly correlated with enrichment factors found in Arenicola marina,
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even though the bioavailable metal concentrations were lower than those extracted by BCR
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procedure (Rosado et al., 2016). It was suggested enzymatic approach is qualitatively more
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reliable than the chemical method in assessing the metals bioavailability in sediments (Ianni et al.,
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2010). Metals in F2 were bound to amorphous Fe and Mn oxides and hydroxides (Nemati et al.,
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2011), which are still considered to be labile and are likely released upon dissolution of Fe/Mn
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oxides under reducing conditions (Rodriguez et al., 2009). This suggests potentially adverse
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effects of both Cu and Cr on the overlying water quality and aquatic organisms. Trace metals
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extracted from F3 were bound mainly to both natural and anthropogenic organic matter by
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complexation and chelation due to their affinity for ligands or active functional groups (Kyziol et
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al., 2006; Nemati et al., 2011), but the decomposition of organic matter into inorganic substances
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by mineralization in an oxidizing environment would induce the release of organically bound
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metals into the overlying water body (Zhuang et al., 2016; Zhang et al., 2017). Whereas, at least
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one-third of trace metal contents were not released, because these metals in residual fraction were
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retained within the stable crystal structures of primary and secondary minerals (Rodriguez et al.,
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2009; Nemati et al., 2011).
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In terms of the spatial distribution of the geochemical fractions of trace metals in sediments, the
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proportions of the non-residual fractions (F1 + F2 + F3) were higher in the middle and upper
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reaches (cores S1−S4) than in the lower reach (S5), suggesting an anthropogenic input (Zhuang et
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al., 2016). This was verified by our previous report that both Dongguan and Shenzhen City
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contributed significantly to the trace metal concentrations of river water in the upper and middle
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reaches (Gao et al., 2016). F4, accounting for 75.6%, 80.7%, 69.0%, 78.4%, and 75.1% of the
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total concentrations of Cd, Cr, Cu, Ni, and Zn, respectively, was predominant in core S5. This core
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could be regarded as a background sample, with trace metals derived mainly from parent minerals.
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In the vertical direction, F1 decreased with depth in cores S1−S4, indicating a sustained increase
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in the input of trace metals through anthropogenic activities in recent years. As reported by the
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previous study that the discharge of industrial effluents and domestic sewage promoted
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sedimentation rate of the Shima River over the last two decades, thus resulting in a significant
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increase in excess metal flux in the upper parts of sediment cores (Gao et al., 2017). Furthermore,
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the sediment Fe/Mn ratio decreased substantially from a depth of 22 cm to the top sediment layer
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in core S4, likely implying a shift in the redox status of sediment to a stronger oxidizing
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environment (Gao et al., 2017). This was responsible for the accumulation of trace metals in the
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surface sediment due to the formation of Fe and Mn oxides, which scavenged a significant amount
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of exchangeable metals. Similarly, a distinct decreasing tendency with depth was observed in F2
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in sediment cores, especially at sites S3 and S4. This could be mainly attributed to two reasons.
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On the one hand, the relatively lower excess metal flux contributed less to metal concentrations in
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sediment; on the other hand, the relatively higher OC contents (mean: 33.6 g·kg-1 for S3 and 39.6
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g·kg-1 for S4) could markedly enhance the reductive dissolution of Fe and Mn compounds
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(Mukwaturi and Lin, 2015), then reducing F2 fraction in the lower part of the sediment cores. It is
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worth noting that the sediment environment of the Shima River is slightly acidic, with a mean pH
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of 6.24 (range, 4.90–7.30). This provided ideal conditions for the release of acid-soluble trace
325
metals from sediment into the overlying river water, indicating that the river sediment could act as
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a secondary pollution source for trace metals, especially for Cd, Ni, and Zn.
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Fig. 2 Spatial and vertical distributions of the total concentrations and geochemical fractions of trace metals in
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sediment cores.
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3.3 Enrichment of trace metals in sediment
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The EF is widely used for quantitative determination of the pollution status of hazardous
333
elements (e.g., trace metals) in soil/sediment (Varol and Şen, 2012; Hu et al., 2013). As shown in
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Fig. 3, the EF values ranged from 7.10 to 27.1 for Cd, 0.161 to 7.40 for Cr, 1.16 to 7.04 for Cu,
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1.30 to 14.0 for Ni, and 1.47 to 28.7 for Zn. The mean EF values decreased in the order of Cd
336
(15.1) > Zn (9.45) > Cu (7.04) > Ni (5.38) > Cr (2.14), indicating severe enrichment of Cd, minor
337
enrichment of Cr, and moderately severe enrichment of Zn, Cu, and Ni. Significant positive
338
correlations (p <0.01) between the concentrations of non-residual metal fractions (F1 + F2 + F3)
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and their corresponding EF values were found (Fig. 3). This further verified that anthropogenic
340
inputs, related mainly to industrial effluent and domestic sewage, contributed significantly to the
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enrichment of trace metals in the sediment of the Shima River. Because bioavailability and
342
toxicity are dependent on the geochemical form and concentration of trace metals in sediments,
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trace metals with high EF values and large labile fractions have been reported to be more mobile
344
and toxic to aquatic organisms (Islam et al., 2015). Therefore, Cd and Zn, with the highest EF
345
values, were significantly enriched in F1 and might pose a substantial hazard to the ecology of the
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Shima River.
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Fig. 3 Relationships between the concentrations of the non-residual metal fractions (F1 + F2 + F3) and their
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corresponding enrichment factors (EFs)
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3.4 Acute toxicity of sediment
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A LUMISTox® bioassay using V. fischeri was conducted in the liquid phase of resuspended
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sediments from the Shima River. As shown in Fig. 4, the acute toxic effects of sediment exhibited
354
clear spatial differences, and the mean TUa values of S3 (0.972) and S4 (0.973) were significantly
355
higher than those of S1 (0.601), S2 (0.677), and S5 (0.396) (p <0.05). TUa values exceeding 0.4
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were considered to represent slight acute toxicity in cores S1, S2, S3, and S4, whereas no acute
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toxicity was found in core S5. As a result of severe pollution caused by Cu, Ni, and Zn, the LIR of
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V. fischeri decreased from 38.4% at S1 to < 10% at S5 in the river water, representing slight acute
359
toxicity in the upper and middle reaches, but little toxicity in the lower reach (Gao et al., 2015).
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The spatial distribution of toxic effects in the river water was roughly in accordance with that in
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the sediment. This implied that aquatic/benthic organisms exposed to the upper and middle
362
segments of the Shima River would experience acute toxic effects. It was obvious that the profiles
363
of TUa, which decreased with depth in all sediment cores except S5, were consistent with the
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vertical distribution of the non-residual forms of trace metals (Fig. 2), indicating a close
365
relationship between toxic effects and the labile fraction of trace metals in sediments.
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Fig. 4 Profiles of the acute toxicity units (TUa) of sediment cores from the upper to lower reaches of the Shima
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River (abc, significant differences in TUa in sediment cores at p <0.05).
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To provide detailed information on toxicity levels in sediments from the Shima River, the TUa
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per gram of sediment (TUa/g) in the study area was compared with those reported at other sites
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(Table 2). Significantly higher TUa values, with a range of 0.3−8700, were found in European
373
harbor, coastal, and estuarine sediments (Onorati and Mecozzi, 2004; Libralato et al., 2008; Usero
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et al., 2008; Serafim et al., 2013; Rosado et al., 2016; Usero et al., 2016), whereas TUa values <1
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from Thessaloniki Bay (Zabetoglou et al., 2002) were comparable with those found in the present
376
study (0.269−1.84). It was surprising that the sediment TUa values displayed such clear
377
differences, even though the total concentrations of trace metals and F1 in sediments from the
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Marghera industrial port and Huelva Estuary were comparable with or lower than those in the
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Shima River, probably indicating different assay procedures. However, the sediment dry weight
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and extract (2% NaCl) volume did not seem to be the key factors affecting the toxic effects, as
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their ratios were within a narrow range (1:4−1:10). Therefore, filtration of the liquid phase from
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the resuspended sediment was likely the main cause of the relatively low LIR in the present study.
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Ringwood et al. (1997) reported that the density of bacterial cells was drastically decreased after
384
adding V. fischeri to a liquid phase containing an abundance of silt-clay particles (< 0.63 µm); this
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decrease was due to the absorption of bacteria onto the particles, resulting in a loss of
386
luminescence intensity in the liquid phase and therefore indicating a higher toxicity than expected
387
(Parvez et al., 2006). An identical bioassay procedure conducted on sediments collected from
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Thessaloniki Bay and Shima River indicated less toxic effects, which might be more realistic and
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rational regarding the weak ion exchange capacity of NaCl solution with solid metals.
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Table 2 Toxicity of sediments from the Shima River and other areas worldwide.
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3.5 Toxicity assessment based on chemical models
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The ∑TU was used to assess the toxic effects of trace metals in the sediments by enabling
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comparisons of the potential toxicities among various sediments based on chemical models
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(Zhang et al., 2016). Figure 5 shows the TU values of individual metals and the ∑TU values of
397
different sediments. The distribution of the ∑TU profiles was similar to that of the TUa profiles
398
from the bioassay (Fig. 4). ∑TU was generally decreased with depth in cores S1−S4, while the
399
∑TU profile remained constant in core S5. The mean ∑TU values of the five sediment cores
400
followed a descending order of S3 (11.0) > S4 (9.61) > S1 (3.70) > S2 (3.68) > S5 (1.49).
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Pederson et al. (1998) assessed sediment toxicity in the amphipod C. volutator and reported that
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no significant mortality was likely to occur at ∑TU < 4. In the present study, sediment with a ∑TU
403
< 4 (e.g., below 5 cm in core S1 and below 4 cm in cores S2 and S5) did not present an obvious
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acute toxicity to V. fischeri (Fig. 4), whereas more significant toxicity was found in cores S3 and
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S4, with ∑TU values ranging from 6.46 to 18.6 and from 5.20 to 21.4, respectively. This further
406
indicated that a ∑TU of 4 could be an important threshold to identify toxic effects associated with
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trace metals in the sediment.
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The TU due to an individual metal followed a descending order of Ni (1.87) > Zn (1.84) > Cu
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(1.26) > Cr (0.722) > Cd (0.207), and the mean contributions of each trace metal to the ∑TU were
410
35.2 ± 7.82% for Ni, 28.1 ± 8.47% for Zn, 18.3 ± 6.08% for Cu, 13.7 ± 5.22% for Cr, and 4.79 ±
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2.06% for Cd, demonstrating that Ni and Zn accounted for the majority of the overall sediment
412
toxicity. Cd contributed the least to the ∑TU, despite its high enrichment in sediment based on EF
413
values (Fig. 3). An assessment using the ∑TU-based approach could underestimate the potential
414
hazards of Cd because of the relatively high PEL value of Cd (Zhang et al., 2016). Cd is one of the
415
most hazardous elements to organisms, with a high toxic response factor in a potential ecological
416
risk index (Håkanson, 1980). Therefore, it was suggested that a range of assessment systems
417
related to ecological risk and/or potential toxicity should be conducted for comprehensive and
418
accurate determination of the practical hazards of various trace metals on the environment.
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Fig. 5 Profiles of the sum of all toxicity unit (∑TU) and toxic risk index (TRI) values based on the total
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concentration of trace metals at different sampling sites
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According to the SQGs proposed by MacDonald (1994), the TEL is considered reliable if the
424
adverse effects are less than 10% within the minimal effect range, while the PEL is considered
425
reliable if the adverse effects exceed 65% of the probable effect range. Therefore, aquatic
426
organisms exposed to the TEL of trace metals in sediment would exhibit only limited acute toxic
427
effects; however, the possibility of chronic toxic effects cannot be dismissed if the exposure
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duration is long enough. Therefore, a TRI integrating the TEL and PEL was applied to assess toxic
429
risks in terms of both acute and chronic toxic effects in aquatic organisms exposed to sediment
430
contaminated by trace metals. As shown in Fig. 5, the spatial and vertical distributions of the TRI
431
values were similar to those of ∑TU in all sediment cores. The mean TRI values of the five
432
sediment cores followed a decreasing order of S3 (27.4) > S4 (25.0) > S2 (8.89) > S1 (8.54) > S5
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(3.76), indicating very toxic risks (TRI > 20) in cores S3 and S4. Additionally, cores S1 and S2
434
presented moderate toxic risks, while little toxic risk was associated with core S5. The toxicities of
435
cores S1−S4 were higher when determined based on the TRI values compared with the ∑TU
436
values, which was not surprising, because chronic toxic effects are taken into consideration when
437
using the TRI. The mean TRI values of individual metals followed a decreasing order of Cu
438
(5.22) > Ni (3.79) > Zn (3.13) > Cr (1.65) > Cd (0.916), with mean contributions of 30.3 ± 8.63%,
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29.1 ± 7.19%, 19.5 ± 6.58%, 12.6 ± 4.60%, and 8.56 ± 3.58%, respectively. The significant
440
contribution of Cu to the TRI was ascribed mainly to its relatively low TEL. Fernandes et al.
441
(2007) found that Cu concentrations of 83 ± 52 mg·kg-1 in sediment induced significant
442
bioaccumulation in the livers of leaping grey mullet Liza saliens after long-term exposure,
443
yielding a bioaccumulation factor of 305%. Furthermore, the Cu and Zn loadings of bivalve
444
species (Anodonta sp. and Unio pictorum) and gastropod species (Radix ovata and Viviparus sp.)
445
collected from a site polluted by trace metals exceeded the environmental concentrations
446
(Gundacker, 2000), and the ingestion of sediment was both an important source of Cu and Zn and
447
a cause of toxicity in benthic mollusks and amphipods (Gundacker, 2000; King et al., 2006). This
448
highlighted the potential chronic toxicity of sediment to aquatic organisms in the Shima River.
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3.6 Identification of toxic elements to V. fischeri
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A BMI integrating F1 of the various trace metals was used to assess the total bioavailability of
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metals trapped in sediment. As shown in Fig. 2, core S5 was considered as the background sample
452
in the study area, because Cr, Ni, Cu, and Zn concentrations were close to the local soil
453
background values (Table 1) and all trace metals were mainly in the residual form and derived
454
from minerals; thus, the mean F1 values of Cd, Cr, Cu, Ni, and Zn in core S5 were selected as
455
reference values. Figure 6 shows the distribution of the BMI values in cores S1−S4, with ranges of
456
0.332−18.3 for S1, 2.69−11.8 for S2, 12.0−31.8 for S3, and 12.7−47.0 for S4. The distribution of
457
the BMI values was similar to those of TUa, ∑TU, and TRI, confirming close relationships of the
458
bioavailable fraction of trace metals with the observed and estimated toxic effects.
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Fig. 6 Profiles of the bioavailable metal index (BMI) at sampling sites S1−S4
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As shown in Fig. 7, TUa values from the biological assay were positively related to the BMI,
463
∑TU, and TRI values (p <0.01), with determination coefficients (R2) of 0.619, 0.833, and 0.822,
464
respectively. A significant positive correlation between TUa and BMI suggested that acid-soluble
465
trace metals may be the geochemical fraction mainly responsible for the acute toxicity of the
466
sediment. A high concentration of trace metals in this fraction will lead to more bioavailable trace
467
metals in the sediment, resulting in more severe toxic effects. Rosado et al. (2016) reported a
468
stronger correlation (R2 = 0.704) between acute toxicity and the BMI, confirming that acid-soluble
469
metals obtained from the modified BCR method are an important source of sediment toxicity.
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Fig. 7 Linear fitting of acute toxicity units (TUa) and the bioavailable metal index (BMI), sum of all toxicity units
472
(∑TU), and toxic risk index (TRI) in sediment cores S1−S4.
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In the present study, PCA was used for identification of trace metals causing sediment toxicity.
475
Firstly, to examine the adequacy of the data for PCA, the KMO index and Bartlett's test were used.
476
The calculated KMO value was 0.821 and the significance level of Bartlett's sphericity test was 0,
477
validating the suitability of the dataset for PCA. As shown in table S3, two components with
478
eigenvalues > 1 explained 92.0 % of the total variance. Acid-soluble Cd, Cr and Cu were clustered
479
in component 2, implying their little effects on the toxicity from bioassay. Whereas, acid-soluble
480
Ni and Zn and TUa with coefficients > 0.7 exhibited strong loadings (Singh et al., 2004) on
481
component 1, likely indicating that acute toxic effects were mainly associated with the
482
acid-soluble fraction of these two metals (Fig. 8). In addition, the acid-soluble Zn and Ni
483
averagely accounted for 62.6 % and 43.2 % of the total concentrations in cores S1−S4, which
484
were obviously higher than those for Cd (37.0%), Cr (2.37%) and Cu (24.3%) (Fig. 2), indicating
485
that of the metals investigated, Zn and Ni were likely responsible for the majority of the acute
486
toxicity in sediments from the Shima River. This could be supported by Gao et al. (2015), who
487
reported that Zn at concentrations of 85 ‒ 494 µg·L-1 in the Shima River was the main toxicant
488
responsible for the observed acute toxic effects on V. fischeri
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Fig. 8 The loadings of acid-soluble metals on different components.
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Conclusion
493
This study identified spatial and vertical variations in the geochemical fractions of trace metals
494
(Cd, Cr, Cu, Ni, and Zn) in sediment cores from the Shima River. The mobility and bioavailability
495
of trace metals followed a decreasing order of Zn > Cd > Ni > Cu > Cr. A positive correlation
496
between non-residual metal fractions (F1 + F2 + F3) and EF values further verified the
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anthropogenic contribution to trace metal pollution in sediments. The LUMISTox® bioassay using
498
V. fischeri indicated a slightly acute toxicity in the upper and middle reaches but few toxic effects
499
in the lower reach, while a relatively higher toxicity was found in the top sediment layers.
500
Sediment toxicity observed in bioassay showed a positive relationship with the estimated toxic
501
effect based on chemical models. Acid-soluble metals obtained from the modified BCR method
502
were found to be mainly responsible for sediment toxicity, and acid-soluble Zn and Ni contributed
503
significantly to acute toxicity in sediments.
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Acknowledgement
The authors are very grateful to reviewers for providing invaluable suggestion and comment.
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This work was financially supported by the National Natural Science Foundation (41701585 and
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41611140112), the Fundamental Research Funds for the Central Universities of China (17lgpy40)
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and the Natural Science Foundation of Guangdong, China (2017A030310309).
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Table 1 Statistical summary of heavy metal concentrations in different sediment cores from Shima River (mg·kg-1)
Cr
Ni
Cu
Zn
50.9±36.2a
51.6±36.9ab
55.9±44.8a
412±367b
S1 (n=15)
13.8–108
14.1–111
12.0–129
63.1–1030
71.5±76.3a
63.1±60.4b
66.1±76.6a
275±372ab
S2 (n=15)
23.0–279
24.1–216
12.2–271
57.6–1310
219±113b
146±33.8d
262±146b
941±446c
S3 (n=15)
93.8–427
105–201
101–512
367–1680
183±109b
116±63.4c
273±147b
789±409c
S4 (n=15)
90.2–469
48.5–253
126–630
379–1700
54.4±5.46a
22.7±2.48a
23.4±5.27a
75.9±10.1a
S5 (n=15)
48.5–69.0
17.7–25.4
18.3–39.1
65.2–93.2
BV
50.5
14.4
17.0
47.3
TEL
52.3
15.9
18.7
124
PEL
160
42.8
108
271
Data expression: mean ± standard deviation (range of heavy metal concentration).
Cd
0.601±0.409a
0.204–1.35
0.585±0.503a
0.229–1.88
1.28±0.273b
0.946–1.75
1.38±0.292b
0.870–1.91
0.506±0.050a
0.447–0.657
0.056
0.68
4.21
Sc
5.99±2.43b
3.38–10.2
4.46±2.59a
2.06–10.8
10.5±0.404c
9.89–11.3
11.1±1.02cd
8.82–12.6
12.1±1.09d
9.14–13.6
8.13
−
−
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Sampling site
a, b, c, and d indicate significant differences (p<0.05).
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BV: Environmental background value for the soil of Guangdong province (China National Environmental Monitoring Centre, 1990)
TEL and PEL: the threshold effect level and the probable effect level (Esen et al., 2010; MacDonald, 1994)
Table 2 Toxicity of sediments from the Shima River and other areas worldwide.
Shima River, China
Thessaloniki Bay,
Greece
Marghera industrial
port, Italy
Leghorn Harbor, Italy
Sediment (g): 2%
NaCl solution (mL)
1.5:10
10:40
10:100
7:35
Solid-phase test object
TUa/g
Source
0.269−1.84
This study
<1
Zabetoglou et al., 2002
54.8−330
Libralato et al., 2008
98−844
Onorati and Mecozzi, 2004
54−8700
Rosado et al., 2016
19−73
Usero et al., 2008
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Study area
Liquid phase from the resuspended
sediment after filtration through a
0.45 µm polypropylene membrane
Liquid phase from the resuspended
Huelva Estuary, Spain
3:30
Atlantic coast, Spain
N.A.
Cadiz Bay, Spain
3:30
39−240
Usero et al., 2016
Portuguese estuaries
7:35
0.3−4.9
Serafim et al., 2013
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D
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N.A.: not available.
sediment without filtration
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Fig. 1 Location of the study area and sampling sites.
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S1
0
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
Depth (cm)
350
0
150
0
300
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
150
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EP
1
2
3
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6
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8
9
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Depth (cm)
0
500 1000 1500 2000
1
2
3
4
5
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8
9
10
11
12
13
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S4
0
500
250
350
0
150
0
0
1
2
3
4
5
6
7
8
9
10
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12
13
14
15
350
150
S5
0
500
500
Cu (mg kg-1)
700
0
350
700
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
Ni (mg kg-1)
0
300
Zn (mg kg-1)
0
2
4
6
8
10
12
14
16
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24
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30
250
1
2
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150
300
1
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2
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24
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30
500 1000 1500 2000
2
Cr (mg kg-1)
Ni (mg kg-1)
300
Zn (mg kg-1)
0
1
1
2
3
4
5
6
7
8
9
10
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12
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2
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1
2
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7
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9
10
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12
13
14
15
500 1000 1500 2000
0
Cu (mg kg-1)
700
Ni (mg kg-1)
300
Zn (mg kg-1)
Zn (mg kg-1)
0
250
Cd (mg kg-1)
2
Cr (mg kg-1)
Cu (mg kg-1)
Ni (mg kg-1)
Ni (mg kg-1)
1
2
4
6
8
10
12
14
16
18
20
22
24
26
28
30
2
4
6
8
10
12
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16
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20
22
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28
30
0
700
1
2
3
4
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6
7
8
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10
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14
15
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
0
350
0
S3
0
Cu (mg kg-1)
700
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
1
2
3
4
5
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14
15
500
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
Cu (mg kg-1)
Depth (cm)
250
Cd (mg kg-1)
2
Cr (mg kg-1)
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
0
1
2
3
4
5
6
7
8
9
10
11
12
13
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15
S2
0
500
1
2
3
4
5
6
7
8
9
10
11
12
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15
1
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Depth (cm)
250
0
2
Cr (mg kg-1)
Cr (mg kg-1)
0
1
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TE
D
Depth (cm)
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
Cd (mg kg-1)
Cd (mg kg-1)
1
SC
Cd (mg kg-1)
0
Zn (mg kg-1)
0
500 1000 1500 2000
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
500 1000 1500 2000
Residual
Oxidizable
Reducible
Acid-soluble
Fig. 2 Spatial and vertical distributions of the total concentrations and geochemical fractions of trace metals in
sediment cores.
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Fig. 3 Relationships between the concentrations of the non-residual metal fractions (F1 + F2 + F3) and their
corresponding enrichment factors (EFs)
TUa
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
0.4
0.8
1.2
2
0
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
S3
TUa
S2
1.6
0.4
0.8 1.2
1.6
a
2
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
0
S4
TUa
0.4
0.8 1.2
1.6
2
TE
D
Depth (cm)
0
EP
S1
a
b
0
2
4
6
8
10
12
14
16
18
20
22
24
26
28
30
S5
TUa
0.4
0.8
1.2
1.6
0
2
b
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
TUa
0.4
0.8 1.2
1.6
2
a
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Fig. 4 Profiles of the acute toxicity units (TUa) of sediment cores from the upper to lower reaches of the Shima
River (ab, significant differences in TUa in sediment cores at p <0.05).
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∑TU
0
12 16 20 24
1
2
3
4
5
6
7
8
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S1
4
Depth (cm)
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8
0
0
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4
6
8
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28
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0
1
2
3
4
5
6
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S2 15
TRI
0
10 20 30 40 50 60
2
4
6
8
10
12
14
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22
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28
S3 30
∑TU
12 16 20 24
S4
TRI
10 20 30 40 50 60
8
S3 2
2
3
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5
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12
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15
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2
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S1 15
∑TU
12 16 20 24
S2 1
TRI
10 20 30 40 50 60
1
2
3
4
5
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8
9
10
11
12
13
14
15
∑TU
12 16 20 24
1
2
3
4
5
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7
8
9
10
11
12
13
14
15
TRI
0
8
0
4
8
12 16 20 24
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
S5
Cd
Cr
Cu
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Ni
Zn
TRI
0
10 20 30 40 50 60
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4
1
2
3
4
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6
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9
10
11
12
13
14
15
SC
∑TU
0
S4
10 20 30 40 50 60
Cd
Cr
Cu
Ni
Zn
S5
Fig. 5 Profiles of the sum of all toxicity unit (∑TU) and toxic risk index (TRI) values based on the total
concentration of trace metals at different sampling sites
S1
0
S2
0
10 20 30 40 50
1
2
3
4
5
6
7
8
9
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15
10 20 30 40 50
EP
Depth (cm)
1
2
3
4
5
6
7
8
9
10
11
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13
14
15
0
10 20 30 40 50
BMI
BMI
BMI
BMI
0
S3
10 20 30 40 50
2
4
6
8
10
12
14
16
18
20
22
24
26
28
30
S4
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Fig. 6 Profiles of the bioavailable metal index (BMI) at sampling sites S1−S4
Fig. 7 Linear fitting of acute toxicity units (TUa) and the bioavailable metal index (BMI), sum of all toxicity units
(∑TU), and toxic risk index (TRI) in sediment cores S1−S4.
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1
Cu Cr
Component 1
Cd
Component 2
0.5
Zn
Ni
TUa
0
-1
-0.5
0
0.5
1
-1
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Fig. 8 Loadings of acid-soluble metals on different components.
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Metals cause sediment toxicity at the upper and middle reach sites of the Shima river
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∑TU = 4 is a critical threshold to identify toxic effects related to sediment metals
Acid-soluble metal is the chemical fraction mainly responsible for sediment toxicity
Acid-soluble Zn and Ni contribute significantly to acute toxicity in sediments
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