J Soils Sediments (2010) 10:65–76 DOI 10.1007/s11368-009-0156-z
INTERCOMPARTMENT • RESEARCH ARTICLE
Physicochemical and ecotoxicological evaluation of estuarine water quality during a dredging operation Sandro R. Urban & Albertina X. R. Corrêa & Carlos A. F. Schettini & Paulo R. Schwingel & Rafael M. Sperb & Claudemir M. Radetski
Received: 21 February 2009 / Accepted: 23 October 2009 / Published online: 13 November 2009 # Springer-Verlag 2009
Abstract Purpose Most of the information concerning the effects of contaminated sediments on estuarine organisms deals with the impacts of bed forming sediments. The ecotoxicological potential at the time of a dredging operation is more difficult to assess, and few studies have dealt specifically with resuspended contaminated sediments. The aim of this study was to determine whether release of contaminants through sediment resuspension during a dredging operation in the Itajaí-açu estuary (Brazil) changed the water quality classification and had an ecotoxicological impact on the near-field water column during the critical moment of this operation. Materials and methods Waters from two sites (control and dredged sites) were analyzed for physicochemical parameters before, during, and after a dredging operation. In parallel, a short-term, sensitive battery of biotests (bacteria, algae, and daphnids) was performed with water samples before and during this operation according to the ISO bioassay protocols.
Responsible editor: Susanne Heise. S. R. Urban : A. X. R. Corrêa : C. A. F. Schettini : P. R. Schwingel : R. M. Sperb : C. M. Radetski (*) Centro de Ciências Tecnológicas da Terra e do Mar, Universidade do Vale do Itajaí, Rua Uruguai, 458, Itajaí, Santa Catarina 88302-202, Brazil e-mail:
[email protected] Present Address: C. A. F. Schettini Instituto de Ciências do Mar–LABOMAR, Universidade Federal do Ceará–UFC, Fotaleza, Ceará, Brazil
Results and discussion No short-term toxicity was observed with waters collected before or during the dredging operation. The results showed that desorption of contaminants from suspended particles of sediments with a low level of contamination during a dredging operation lowered the water quality in the near-field water column but that this did not promote significant acute toxicity effects on the organisms tested. Conclusions More detailed studies are needed (e.g., the question of the reliability of biotests under turbulent, particle-rich conditions) to fully understand this complex issue regarding water column ecotoxicity during the whole dredging operation and to support decisions on the management of dredging activities. Keywords Dredging . Ecotoxicology . Environmental impact . Estuaries . Metals . PAHs . Water quality
1 Introduction In coastal areas, estuaries have been recognized as being of particular ecological and economic importance (USEPA 1999, 2000), but waters and sediments in estuarine areas are subject to multiple anthropogenic or naturally occurring stress factors such as physical, chemical, and microbiological agents originating from agricultural and urban runoff, municipal sewage, industrial wastewater, and navigation traffic (Rörig 2005; Darbra et al. 2009; Viers et al. 2009). In terms of their physicochemical characteristics, estuaries are not steady-state systems, which make the spatial and temporal assessment of their quality an interesting challenge (Förstner 2004). Additional complexity derives from sediment dredging operations (Gustavson et al. 2008),
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which are necessary to open new harbor terminals or for the maintenance of navigable waterways. This operation causes sediment resuspension, which is defined by Hayes and Engler (1986) as sediment particles suspended in the water column during the dredging operation that do not rapidly settle out of the water column. According to Collins (1995), regardless of the type of dredging operation, there are three sediment features that influence the magnitude and distribution of resuspended sediment in the near-field water column: (a) the physical character of the sediments being dredged (quantified by grain size and distribution and specific gravity—relative to the overlying waters); (b) the condition of the in situ sediments as reflected by in situ bulk density, void ratio, and other similar physical measurements; and (c) the physicochemical characteristics of the sediment or the overlying waters (e.g., salinity), which might affect the cohesiveness and consequently the flocculation and settling of sediment particles. Eggleton and Thomas (2004) present a broad review of the factors affecting the release and bioavailability of contaminants during sediment disturbance events and recently, a German initiative started the Floodsearch project, which aims to combine methodologies of hydraulic engineering and ecotoxicology in a new interdisciplinary approach to assess the risks associated with the remobilization of particulate-bound contaminants often observed after severe flood events (Wölz et al. 2009). Although the contaminant contents in sediments are site and time dependent, generally, metals such as copper, lead, mercury, or zinc and organic compounds such as pesticides, PCBs, and PAHs are the major contaminant constituents (Long 2002). The physical and biological impacts of dredging operations, which are often related to postdredging activities, are not necessarily directly related to the chemical contamination levels (Wilber and Clarke 2001). Most of the information concerning the effects of contaminated sediments on estuarine organisms deals with the impacts of bed forming sediments. The ecotoxicological potential of dredged contaminated sediments is more difficult to assess and few studies have dealt specifically with resuspended contaminated sediments (Bonnet et al. 2000). Thus, the purpose of this study was to assess the water quality and potential ecotoxicological effects of contaminants resolubilized through sediment resuspension as a result of a mechanical dredging. To achieve our goals, physicochemical analysis of water was carried out before and during a dredging operation. Also, a battery of biotests composed of test species belonging to the three trophic levels of aquatic food chains, i.e., producers (algae), primary consumers (daphnids), and decomposers (bacteria), was performed to assess the ecotoxicological
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profile of estuarine waters before and during a dredging operation.
2 Material and methods 2.1 Estuary characteristics The rio Itajaí-açu estuary is located in the south of Brazil at 26.9°S and 48.66°W in the state of Santa Catarina. The estuary comprises the terminal portion of the rio Itajaíaçu, which drains a basin area of 15,500 km2 encompassing 47 municipalities, including a major industrial area in southern Brazil. The estuary flows over a coastal plain and its morphology resembles a deltaic front estuary according to Fairbridge’s physiographic classification (Fairbridge 1980). The estuarine width is almost constant, around 200 m. The bathymetry ranges from 5 to 9 m with absence of sand banks, shallows, and tidal flats. The average annual discharge of the rio Itajaí-açu is 228 m3 s−1, with historical minimum and maximum values of 17 and 5,390 m3 s−1, respectively (Schettini 2002). The river discharge is low most of the time, under 150 m3 s−1, with sparse discharge peaks produced by rain events in the basin. Significant changes in the estuarine structure can be observed when the discharge flow is higher than 500 m3 s−1 (Schettini 2002). The regional tidal regime is microtidal semidiurnal, ranging from 0.4 to 1.2 m during neap and spring tide periods, respectively (Schettini 2002). The physical setting of (a) a small tidal range, (b) a highly variable river discharge regime, and (c) a deep and uniform channel morphology results in a highly stratified estuarine structure (Schettini et al. 2005). A two-layered structure separated by a halocline is observed most of the time, being stronger during neap tide periods. The salt intrusion during low discharge periods extends up to 30 km from the mouth, and all salt is flushed out when the river discharge exceeds around 1,000 m 3 s −1. The former situation usually lasts from a few weeks to a couple of months; while the latter usually lasts from a few hours to a few days. The dredging operation used as the “field laboratory” for this assessment was performed at nearly 8 km from the estuary mouth, comprising an extension of nearly 250 m of margin and 30 m width. Although the volume of dredged material is relatively small (37,500 m3), it is one of several dredging operations along the Itajaí-açu estuary. The operation lasted for approximately 2 months, and the goals of the dredging were to increase the water depth near a pier for ship mooring and to raise the land level on an adjacent land area using the dredged material. The dredger used was a small hopper dredge with a storage capacity of less than 300 m3.
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2.2 Collection, storage, and manipulation of water and sediments Three collection events at each of the two sites studied (upper estuary 26°51′56.37″ S 48°41′26.40″ W and dredging zone 26°52′42.86″ S 48°41′25.89″ W) were carried out consecutively for 3 days, and the grab water sample from each event was collected with a Van Dorn bottle (adapted to high volume) sampler, while sediments were collected with a small Van Veen grab sampler. The Van Dorn bottle was attached to the dredger apparatus, which ensure sampling near the maximum turbidity area during dredging operation. The Itajaí-açu estuary is affected mainly by semidiurnal tides, and samples were taken each day at the low water phase, at the same time, with similar meteorological conditions. The distance between the different collection points was approximately 80 m. The upper estuary site (number 1) was used as the control, and the dredging site (number 2) was the disturbed site. It is important to note that water was sampled when the turbidity was at the maximal level (during the dredging operation). Water and sediment samples were stored in a refrigerator (−18°C) in sealed, completely filled, polyethylene buckets until manipulation. To carry out the bioassays, a composite sample (approximately 50:50% mixture) of deeper and superficial water was filtered (Millipore FG filters, 0.45µm pore size) prior to toxicological testing. All procedures of collection, storage, and manipulation for the toxicological analysis were based on USEPA methodologies (USEPA 2001). 2.3 Suspended sediments The distribution of the water column suspended sediment concentrations was assessed in situ before, during, and after the dredging operation. The probe was lowered into the water column performing readings at 0.5 m intervals. An equation to convert the probe readings to suspended sediment concentrations was obtained through pairs of probe readings, and values of suspended sediment concentration were determined through analysis of water samples. Water samples (surface and deeper water) were taken to quantify the concentration of suspended sediments by a gravimetric method: filtration of a known volume (approximately 20 l) through a preweighed filter. The water and (re)suspended solids collected from this filtration provide sufficient samples for all parameters investigated (USEPA 1998). 2.4 Chemical analysis of water and sediment samples Water samples were filtered prior to chemical analysis (Millipore FG filters, 0.45µm pore size). For the sediment analysis, approximately 30 g of each subsample for PAH
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analysis was spiked with the USEPA cocktail of 16 standard PAHs as internal standards and homogenized with anhydrous sodium sulfate dried at 150°C for 12 h using a mortar and pestle until a free floating powder was obtained. The same procedure was used for the PCB analysis, where another sediment subsample was spiked with congener numbers 28, 52, 101, 118, 138, 153, and 180. Each homogenate was shaken with 60 ml carbon disulfide as the extractor solvent (analytical grade) in a tightly closed brown glass bottle (250 ml) on a mechanical shaker overnight. The supernatant was decanted, and the extraction was repeated successively with 2×30 ml carbon disulfide. All fractions were combined and concentrated to approximately 2 ml using rotary evaporation at a pressure of 760 mmHg and clean-up with column chromatography. A Varian GC 3400 instrument equipped with a DB-WAX column (60 m×0.25 mm, i.d., 0.25µm film thickness) was used to perform the analysis according to the USEPA protocols (USEPA 1996). The injector temperature was 280°C (splitless). Chromatographic conditions included an initial oven temperature of 50°C, with a 2 h 50 min isotherm and a program rate of 7°C/min and a final oven temperature of 240°C with an isotherm of 8 min. The gas carrier was N2, with a column flow of 1 ml min−1, and detection was based on flame ionization. Six solutions (2, 5, 10, 20, 50, and 100 ng ml−1) of the 16 USEPA PAH standards were used for the calibration of the equipment. The quantification was performed by the external standard method and detection limits were 2.0 ng l−1 for water and 0.2 ng g−1 dw for sediments, with spiked recoveries of PAHs between 77– 113%. All of the samples taken were analyzed in triplicate, and the relative standard deviation was less than 20%. Determination of chemical and biological oxygen demand (COD and BOD), nutrients, and metals, as well as colimetric analysis were carried out according to a standardized method (APHA et al. 1995). For metal (and As) determination, approximately 30 g of each subsample was extracted in hot acid conditions (HNO3 +HF+HCl). For As and Hg, hydride generation and the cold vapor techniques were used, and both elements were quantified using atomic absorption spectrometry. The concentrations of Cd, Pb, Cu, Ni, Zn, and Cr were determined using flame or furnace atomic absorption spectrometry, depending on the metal content. For the analytical quality control of metals in total sediments, the reagent blank and the international standard reference materials (US National Institute of Standards and Technology, SRM 1646a) were tested before analysis, and the detection lines for the samples were: Cd (0.001), Pb (0.065), Cu (0.030), Cr (0.027), Ni (0.022), and Zn (0.040) in milligrams per kilogram. The mean certified values for Cd, Pb, Cu, Cr, Ni, and Zn were 0.19, 18.7, 10.72, 27.96, 29.6, and 45.0. The mean recovery values
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(with standard deviation in parenthesis) for the reference material were 0.15 (0.02) for Cd, 16.1 (1.3) for Pb, 10.01 (0.98) for Cu, 22.9 (1.5) for Cr, 27.0 (2.0) for Ni, and 49.0 (2.1) for Zn. Water trace metals were determined from calibration curves using known standards (CASS-4, National Research Council of Canada), and the detection limits were defined as the concentration value, which numerically equals three times the standard deviation of ten replicate blank measurements as follows: 3µg l−1 for Cd, 4µg l−1 for Cu, 5µg l−1 for Zn and Pb, and 20µg l−1 for Ni and Cr. All analyses were carried out on replicate samples, and the coefficient of variation was lower than 15%. 2.5 Ecotoxicity tests 2.5.1 Lumistox test The bacterial (Vibrio fischeri) luminescence inhibition (i.e., Lumistox, Dr. Bruno Lange, Düsseldorf, Germany) test was conducted according to ISO 11384-3 (1996) guidelines at 15±1°C on filtered water samples with salinity adjustment to 35 parts per thousand at pH7. The exposure time was 30 min. The lyophilized bacterial reagent was obtained from Deutsche Sammlung von Mikroorganismen und Zellkulturen (DSM number 7151, Braunschweig, Germany). Each of the five sample dilutions (and the control) was performed in triplicate. 2.5.2 Algal growth inhibition test The algal species used was Skeletonema costatum. Algal tests for each water sample were conducted according to ISO 8692 (1989a) guidelines with three replicates per concentration (or control) of filtered water samples with salinity adjustment to 31 parts per thousand at pH8.1. Potassium dichromate was used as a positive control. The cell density of the mixture was adjusted to 10,000 cells per milliliter by dilution with ISO freshwater algal test medium. The test flasks were incubated on a shaker (100 rpm) with continuous illumination of 70 mE m−2 s−1 (cool-white fluorescent lamps) at 23±2°C. After 72 h of incubation, the inhibitory effect based on fluorescent activity was measured at 685 nm with a Shimadzu RF-551 (Kyoto, Japan) spectrofluorimeter. 2.5.3 Daphnia magna immobilization test The 48-h immobilization test with D. magna was performed in accordance with the ISO 6341 (1989b) standard at 25±2°C using 20 individuals per replicate (less than 24 h old) in 50 ml glass beakers with 40 ml of test medium. Three replicates per dilution were performed for each water
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sample in order to evaluate the variability of the procedure. Each test consisted of five river water dilutions and a control group with salinity adjustment to six parts per thousand at pH7.8. Potassium dichromate was used as a positive control. 2.6 Statistics Statistical analysis was carried out on a microcomputer using TOXSTAT 3.0 software. Responses were presented as the mean (X) of assessed endpoints with the coefficient of variation (CV), which was calculated by dividing the standard deviation by the mean value of the response, multiplied by 100. The William’s test (α≤0.05) was used to obtain the lowest observed effect concentration (LOEC) after applying Shapiro–Wilk’s test for normality and Hartley’s test for homogeneity of variance. For daphnids, Fisher’s exact test was used (α≤0.05).
3 Results and discussion 3.1 Suspended sediment distribution Regarding the objective of this assessment, it is important to remark that the suspended sediment concentration can vary from tens to thousands of milligrams per liter in response to river discharge variations (Schettini and Carvalho 1998). On the other hand, such concentration variations are also observed in response to dredging activities. Tables 1 and 4 show the results of measured suspended particulate matter before and during the dredging operation. In terms of time scale, river discharge peaks last for hours or days when the suspended sediment concentration can rise to several hundreds of milligrams per liter. Dredging operations usually last for several weeks or a couple of months. In terms of spatial scale, river discharge peaks affect the entire estuary, while dredging operation effects are observed locally, in the order of hundreds of meters from the dredged site (Schettini and Carvalho 1998). The high variability of river discharges and high levels of suspended sediment concentration that follow river discharge peaks make it difficult to assess the direct impacts of dredging and evaluate its deleterious effects on the environment. 3.2 Physicochemical and ecotoxicological estuarine aspects before dredging operation Dredging is a physical operation that changes the physical, chemical, and biological composition of the dredged site. During the dredging operation, an amount of sediment is resuspended in the water column allowing contaminants to
J Soils Sediments (2010) 10:65–76 Table 1 Physicochemical and microbiological composition of water at site number 1 (control) and site number 2 in the Itajaí-açu estuary before dredging
Data shown are lowest and highest parameter concentrations sampled at three different times in three different points of the same site (n=9) SPM suspended particulate matter, COD chemical oxygen demand, BOD biochemical oxygen demand, DO dissolved oxygen, PTotal total phosphorus, NKjeldahl Kjeldahl nitrogen, ND not detected a
Fecal coliforms
69 Parameter (unit)
Site 1 (Control)
Site 2
Surface water
Deeper water
Surface water
Deeper water
SPM (mg l−1) Oil and grease (mg l−1) COD (mg l−1) BOD5 (mg l−1) DO (mg l−1) PTotal (%) Nitrate (mg l−1) Nitrite (µg l−1) Nkjeldahl (µg l−1) Cadmium (mg l−1) Copper (mg l−1) Lead (mg l−1)
– 1.6–3.4 22–61 6–21 5.7–7.2 0.13–0.15 3.7–11.2 49–81 33–71 ND ND ND
299.5 ND–2.6 53–93 21–38 4.7–5.9 0.07–0.19 16.2–19.6 23–50 17–79 ND ND–0.050 ND
– 2.0–6.6 38–53 11–17 4.1–4.6 0.02–0.05 5.8–9.0 44–61 27–66 ND ND ND
530.8 ND–3.4 67–94 26–32 3.6–4.4 0.08–0.16 12.1–19.8 20–49 33–83 ND ND–0.073 ND
Nickel (mg l−1) Zinc (mg l−1) Colimetrya (UFC ml−1)
ND ND >25,491
ND ND >25,491
ND ND >25,491
ND–0.05 0.06–0.16 >25,491
become partially resolubilized. According to the literature data, resolubilization is mainly due to the release of pore water containing dissolved chemicals, desorption from sediment particles, and loss of particulate-bound contaminants (Förstner 1987; Förstner and Salomons 2008). Thus, to understand the magnitude of physicochemical changes in the water quality during a dredging operation, it is necessary to know the background values of both the water and sediment at the sites before dredging, which are shown in Tables 1 and 2. Before dredging, water contaminant levels at site number 1 were slightly greater than at site number 2. In general, organic parameters (except oil and grease and nitrite) were higher in deeper than surface waters at both sites, which can be explained by sedimentation process that settled these contaminants in the estuarine region. High organic content in surface and deeper waters may originate from diffuse urban untreated sewage and fishing industry wastewater discharges. Data for the BOD5/COD mean ratios, ranged from 0.31 (site number 1) to 0.36 (site number 2) showing that natural depuration by aquatic microorganisms could be the causative agent of dissolved oxygen water depletion (minimum of 4.7 mg l−1 at site number 1 and 3.6 mg l−1 at site number 2). No Cd, Pb, pesticides, PAHs, or PCB compounds were detected at the two sites. Metal resolubilization from sediments and urban effluent discharges could be the reason for the higher concentrations of Cu, Ni, and Zn in the deeper waters of site number 2 when compared to the surface water concentrations at the same site. Moderate or high chemical concentrations of contaminants in water can cause adverse effects in terms of public health and aquatic biota through direct toxicity and/or bioaccumula-
tion. To avoid these problems, most federal governments have established chemical water quality criteria, which are intended to protect aquatic life and human uses of natural water bodies. Thus, according to the Brazilian water quality standards (CONAMA 2005), water resources are divided according to the preponderant uses into five classes. However, the river analyzed in this study is not classified by the state environmental agency, but Table 1 shows that water from the Itajaí-açu estuary belongs to class 3 (poor quality) due to the metal (Cu and Zn), organic, and microbiological contents. For the metals, the legal limits for class 2 are 8µg l−1 for Cu and 0.12 mg l−1 for Zn, while organic carbon content must be limited to 5 mg l−1. Coliforms were detected at the two sites examined above the limits set by the Brazilian legislation (i.e.,80%
NT
>80%
NT
>80%
NT
>80%
0 0
− −
Composite sample was collected from the deeper and superficial estuarine water (approximately 50:50% mixture). Data presented are mean of the assessed endpoint and percent coefficient of variation (n=3) NT not tested, LOEC lowest observed effect concentration a
Highest water proportion tested without color correction for the algal test
dredging operation. Values of Cu and Zn reached 0.16 and 0.36 mg l−1 for which the maximum permissible limits are 0.008 and 0.12 mg l−1, respectively, while the As concentration reached 0.13 mg l−1 for which the established limit is 0.069 mg l−1. Likewise, the dredging operation increased the dissolved organic matter in the superficial and deeper waters (e.g., there was a 2.5-fold increase in the mean carbon content). These increases could be attributed to the resolubilization caused by the dredging operation. The effects of dredging on the water contaminant concentrations was firstly, due to mechanical disturbance of the bed caused by the material suction and secondly, because of Table 4 Physicochemical and microbiological composition of water at site number 1 (control) and site number 2 (dredged) in the Itajaí-açu estuary during dredging of site number 2
Data shown are lowest and highest parameter concentrations sampled at three different times in three different points of the same site (n=9) SPM suspended particulate matter, COD chemical oxygen demand, BOD biochemical oxygen demand, DO dissolved oxygen, PTotal total phosphorus, NKjeldahl Kjeldahl nitrogen, ND not detected a
Fecal coliforms
Parameter (unit)
the overflow. Once in the water, the mixture of fine sediments and water sinks quickly to the bottom since its density is considerably higher than that of the surrounding water. Part of the material stays close to the bottom and gradually settles forming the bed. Another part is entrained into the surrounding water becoming a true suspension, which will be advected by currents. In this regard, during dredging operations, resolubilization of metals and organic compounds from the sediments, even in highly contaminated areas, has been reported to be minimal (Ludwing and Sherrard 1988; EVS 1997). Nevertheless, other authors have shown that dredging operations
Site 1
Site 2
Surface water
Deeper water
Surface water
Deeper water
SPM (mg l−1) Oil and grease (mg l−1) COD (mg l−1)
184.6 5.5–7.5 22–32
256.9 2.6–3.1 44–55
85,458 6.1–7.2 52–71
109,213 3.8–4.0 109–157
BOD5 (mg l−1) DO (mg l−1) PTotal (%) Nitrate (mg l−1) Nitrite (µg l−1) Nkjeldahl (µg l−1) Arsenic (mg l−1) Cadmium (mg l−1) Copper (mg l−1) Lead (mg l−1) Nickel (mg l−1) Zinc (mg l−1) Colimetrya (UFC ml−1)
4–6 4.6–5.6 0.04–0.06 3.7–5.1 60–80 7.9–12.9 ND ND ND ND ND ND >25,495
12–17 4.0–4.6 0.03–0.05 17.5–19.6 20–30 7.4–8.2 ND ND 0.008–0.060 ND ND 0.06–0.09 >25,495
11–20 3.0–3.8 0.13–0.15 7.8–15.4 40–61 8.9–11.7 0.09 ND 0.01–0.03 0.03–0.05 ND 0.04–0.09 >25,495
35–48 2.5–2.9 0.09–0.29 17.5–26.7 28–58 7.1–10.6 0.13 ND 0.04–0.16 0.04–0.07 0.03–0.07 0.07–0.36 >25,495
72 Table 5 Physicochemical composition of surface and deeper water samples collected from the control and dredging sites and environmental conditions at the time of water sampling for ecotoxicity tests. Data are mean values with CV≤15%
J Soils Sediments (2010) 10:65–76 Parameter (unit)
Site 1 (Control site)
Site 2 (Dredging site)
Surface water
Deeper water
Surface water
Deeper water
23.7 22.0 2.8 6.92 3.4 62.1
23.7 22.1 4.1 7.59 1.5 293.6
23.7 22.2 3.0 7.08 4.4 66.1
23.7 22.2 5.2 7.48 1.6 776.2
Air T (°C) Water T (°C) Salinity (ppt) pH DO (mg l−1) Turbidity (NTU)
may affect the aquatic biota due to chemical and physical changes in the environment (Fredette and French 2004; Cotou et al. 2005; Wilber et al. 2007). The length of time that sediments are resuspended plays a key role in determining the chemical impact on the water column (Tomson et al. 2003) and the vast majority of resuspended sediment settles close to the dredged area within 1 h and only a small fraction takes longer to resettle (Van Oostrum and Vroege 1994). To determine whether contaminant release during the dredging operation can cause acute environmental impact, the same battery of toxicity tests was used previously to assess water toxicity, i.e., before the dredging operation. Some physicochemical characteristics of the water at the time of the collection of samples for use in the battery of biotests during the dredging operation are shown in Table 5. It can be seen in Table 5 that the deeper water has a higher salinity and lower dissolved oxygen content than the surface water, probably, due to the slight saline intrusion and organic matter remobilization resulting from the dredging, with its subsequent biodegradation by microorganisms.
The results of the ecotoxicity tests during the dredging operation are shown in Table 6. Table 6 shows that neither the EC50 values nor the LOEC was observed when the three test organisms were exposed to the filtered estuarine water. In natural waters, the resorption and/or complexation of contaminants by clay particles and organic compounds coming from overflow and the large dilution factor could explain the unavailability of contaminants and consequent lack of toxicity response. Thus, it should be noted that metals (and As) in oxic sediments could be scavenged by the iron/manganese oxyhydroxides and carbonates associated with solid-phase natural organic matter or bound to the mineral particles in suspension. For the deeper anoxic sediments, the oxyhydroxides dissolve, releasing the metals, but these in turn could be captured by sulfides formed by the reduction of sulfate. At the transition zone between oxic and anoxic environments in the sediment, conditions may allow the formation and maintenance of sulfide phases. In this transition zone, if neither oxyhydroxides nor sulfide phases are present, many metals are solubilized; but there is the presence of fine clay and/or organic matter from overflow,
Table 6 Ecotoxicity results of waters from sites number 1 (control) and number 2 (dredged) during dredging operation at site number 2 Organism test and assessed endpoint
Filtered estuarine water (%) Site
Vibrio fischeri luminescence variation
Control Dredged
Skeletonema costatum growth rate
Control Dredged
Daphnia magna lethality
X CV X CV X CV X CV
Control Dredged
Measured endpoint
0.0
6.25
12.5
25.0
50.0
80.0a
100.0
LOEC
3,115 8.6 3,210 7.8 0.944 9.1 0.950 12.8 0 0
3,068 7.2 2,946 11.8 0.914 7.9 0.936 14.6 0 0
3,032 7.9 3,111 12.9 0.888 6.9 0.898 11.2 0 0
3,050 9.9 3,095 11.7 0.999 10.4 1.034 15.1 0 0
3,061 10.2 3,043 8.5 0.936 8.0 1.044 12.4 0 0
3,043 7.5 3,033 10.1 0.868 7.5 1.054 17.7 0 0
NT
>80%
NT
>80%
NT
>80%
NT
>80%
0 0
− −
Composite sample was collected from the deeper and superficial estuarine water (approximately 50:50% mixture). Data presented are mean of assessed endpoint and percent coefficient of variation (n=3) NT not tested, LOEC lowest observed effect concentration a
Highest water proportion tested without color correction for the algal test
J Soils Sediments (2010) 10:65–76 Table 7 Physicochemical composition (dry weight basis) of superficial sediments from different sites of the Itajaí-açu estuary after dredging operation
Data shown are minimum and maximum parameter concentrations for samples collected at three different times in three different points of the same site (n=9) ND not detected
73 Parameter (unit)
Site 1 After dredging
Site 2 After dredging
Organic Carbon (%) Carbonate (%)
2.0–2.7 4.2–5.2
1.0–1.9 4.2–4.5
Nitrogen (%) Phosphorus (%) Anthracene (µg kg−1) Benzo(a)fluoranthene (µg kg−1) Benzo(k)phenanthrene (µg kg−1) Benzo(a)pyrene (µg kg−1) Crisene (µg kg−1) Fluoranthene (µg kg−1) Naphthalene (µg kg−1) Phenanthrene (µg kg−1) Arsenic (mg kg−1) Chromium (mg kg−1) Copper (mg kg−1) Lead (mg kg−1) Mercury (mg kg−1) Nickel (mg kg−1) Zinc (mg kg−1)
0.68–0.72 0.04–0.06 ND ND ND ND ND ND ND ND 4.3–5.0 22.2–36.1 11.8–28.2 12.6–19.5 ND 12.2–16.8 63.5–77.2
0.48–0.56 0.04–0.07 ND ND ND ND ND ND ND ND 2.7–3.9 15.8–22.2 7.8–9.3 9.5–11.5 ND 6.5–9.5 25.3–48.8
which could adsorb free contaminants. Thus, only a small fraction of metal concentrations are resolubilized and bioavailable under normal conditions (EVS 1997; Maddock et al. 2007). For this reason the USEPA (2007) recommend that assessment of the effects of aqueous metals on aquatic organisms be based on dissolved metal concentrations. In relation to metal ecotoxicity based on the literature, 48 h LC50 values for D. magna of 93µg Cu l−1 and 7,290 µg Ni l−1 have been reported (Khangarot and Ray 1989), and water column concentrations for these metals in a dredging zone were found to be between 10.0–160.0 µg Cu l−1 and 40.0–70.0µg Ni l−1 (see Table 1). For V. fischeri, 30 min EC50 values of 7.1µg Cu l−1 and 669 µg Pb l−1 have been reported (Hsieha et al. 2004), and in our study, water column concentrations for these metals in the dredging zone were between 10.0 and 160.0µg Cu l−1 and 30.0–70.0µg Pb l−1 (see Table 1). In the case of S. costatum, 72 h EC50 values of 45µg Cu l−1 and 142 µg Zn l−1 have been found (Walsh et al. 1988), and in our study, water column concentrations for these metals in the dredging zone were between 10.0 and 160.0µg Cu l−1 and 20.0–200.0µg Zn l−1 (see Table 1). Overall, metal water column concentrations could present some toxicity, but it should be mentioned that toxicity values obtained for a metal solution are quite different from those for complex natural mixtures. Furthermore, the presence of nutrients in the water column can promote increases in organism production, counter-balancing toxic effects from dissolved metals (Rosa et al. 2001). In this regard, our results agree
with other toxicity studies where concentrations of metals released during resuspension of moderately contaminated sediments were not sufficient to be acutely toxic (Eggleton and Thomas 2004). Although we used short-term ecotoxicity endpoints, our results can be compared with other experimental designs carried out to assess the long-term effects caused by disposal of dredged material in water. Bonnet et al. (2000) used D. magna and Hydra attenuata survival to assess the environmental impact of two moderately contaminated freshwater suspension sediments under flow-through conditions. After 96 h of exposure, the overlying water of one of the sediment samples showed Table 8 Metal concentrations in the suspended particles at both, control (site number 1), and dredged (site number 2) sites after dredging of site number 2 Parameter (unit)
Site 1
Site 2
Arsenic (mg kg−1) Cadmium (mg kg−1) Chromium (mg kg−1) Copper (mg kg−1)
5.0±0.6 ND 28.1±1.7 26.1±2.0
4.5±0.5 ND 36.6±1.5 30.0±1.6
Lead (mg kg−1) Nickel (mg kg−1) Zinc (mg kg−1)
15.1±3.2 17.2±0.9 82.1±3.3
13.5±0.9 16.5±0.7 73.1±3.4
The results are the mean values and standard deviation of triplicate samples ND not detected
74
some toxicity toward H. attenuata, which was probably due to the presence of ammonia and copper, whereas D. magna did not reveal any toxicity response. This study, therefore, indicates a possible minor impact of dredging on the estuarine water column biota. Thus, the resuspension of sediments with low level contamination resulted in non measurable toxicity at the dredged site, probably due to the metal speciation since several studies have reported that the metal species were bound to strong binding sites of humic compounds/black carbon and/or bound to or trapped in colloidal materials (Förstner and Wittmann 1981; Lu et al. 1996; Klavins et al. 2000). Nevertheless, it has also been reported that very high levels of resuspended sediments and turbidity do have the potential to affect aquatic organisms by changing the community composition of benthic macroinvertebrates where thin-layer dredged material was disposed of or the sediment was subjected to intense clam dredge-fishing (Wilber et al. 2007; Constantino et al. 2009). However, most of these impacts occur at resuspension levels, and durations that are not typically present during small dredging operations or the effects are comparable to the impact of surface waves on the bottom in wavedominated environments. 3.4 Physicochemical estuarine aspects after dredging operation Physical removal of contaminated sediment can cause a negative impact on benthic organisms, but for heavily contaminated sites, we must bear in mind that a dredging operation could have positive benefits such as permanent removal of contaminants from the aquatic system (Voie et al. 2002; Weston et al. 2002). In this regard, when sediment quality is analyzed after dredging (Table 7), concentrations of As, Cr, Pb, Ni, and Zn are clearly lower than those observed before the dredging operation (see Table 2). Data from site number 2 (Table 8) shows that concentrations of Cu, Pb, Ni, and Zn in the suspended particles after the dredging operation are lower than sediment metal concentrations before the dredging operation, but are in the same order of magnitude as concentrations found in the estuarine sediments after dredging operation. Overall, water and sediment concentrations of contaminants at site number 1 were practically the same before, during, and after the dredging operation, while at site number 2, these concentrations were lower after the dredging operation, with decreases of 31.5% for Zn, 43.5% for As, 54.3% for Cr, 55.2% for Pb, 70.8% for Ni, and 76.9% for Cu. As elutriate analysis from particulate material in suspension showed the presence of metals after the dredging operation, it could be concluded that the lack of water toxicity is due to unavailability of these contaminants to the aquatic biota.
J Soils Sediments (2010) 10:65–76
4 Conclusions The results of the chemical analysis showed that the sediments from the Itajaí-açu estuary had a low or moderate level of contamination. The resuspension events from the dredging operation caused a slight resolubilization of contaminants. Although this was sufficient to deteriorate the estuarine water quality, it was not sufficient to cause toxicological impacts when a battery of tests was carried out with different species of aquatic organisms (bacteria, algae, and daphnids). Thus, the battery of biotests indicated that water samples collected from deeper and surface levels at the critical moment of dredging-resuspension (i.e., when resolubilization could occur at the maximal level) did not exert a significant acute toxicity effect. Thus, the sampling strategy applied at the crucial moment of the dredging operation together with the battery of ecotoxicological tests implemented in this study appears to be a useful tool for toxicity assessment of contaminants released from resuspended particles of contaminated sediments during dredging operations. Nevertheless, it must be emphasized that water samples were collected during the least hydrochemically stable situation of the dredging operation, and this moment is not a representative of the broader hydrodynamic process that occurs in the estuary during this operation. Although the combination of bioassays and chemical analysis could aid the improvement of sediment and water toxicity criteria, more detailed studies are needed (e.g., the question of the reliability of biotests under turbulent, particle-rich conditions) to fully understand this complex issue regarding water column ecotoxicity during the whole dredging operation and to support decisions on the management of dredging activities. Acknowledgments The authors greatly acknowledge the fellowship support of CNPq—Brazilian National Council for Scientific and Technological Development (grants 300898/2007-0 for CM Radetski and 306217/2007-4 for CAF Schettini).
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