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Contaminant release from sediments in a coastal wetland
PII: S0043-1354(98)00286-3
Wat. Res. Vol. 33, No. 4, pp. 909±918, 1999 # 1999 Elsevier Science Ltd. All rights reserved Printed in Great Britain 0043-1354/99/$ - see front matter
CONTAMINANT RELEASE FROM SEDIMENTS IN A COASTAL WETLAND M S. S. S. LAU* and L. M. CHU*
Department of Biology, The Chinese University of Hong Kong, Shatin, Hong Kong, P.R. China (First received February 1998; accepted in revised form July 1998) AbstractÐThe Mai Po Marshes Nature Reserve (Hong Kong) which is one of the most important regional wetland habitats, is under increasing pollution pressure, such that the existence and survival of the natural biota, including many endangered species, is threatened. Water quality in the tidal shrimp ponds (locally called gei wais) of the marshes was so poor, that biological impacts arising from the release of contaminants from the underlying sediments were suggested. A study was conducted to (i) investigate the level of contamination of underlying sediments in Gei wai 13, (ii) examine the eects of environmental factors on the behavior of contaminants in the sediments and (iii) study the toxicity of the sediments. Sediment in Gei wai 13 was contaminated with nutrients and heavy metals, of which nitrogen and phosphorus tend to bind to coarser particles, while copper, nickel and zinc to ®ner ones (silts and clays). Mobility of nitrogen and phosphorus from the sediment increased with increasing levels of salinity and temperature of the overlying water. In respect of the drying eect, the solubility of nitrogen and phosphorus of air-dried sediments was lower than that of wet sediments. Sediment extract toxicity tests indicate, that the highly contaminated sediment resulted in a low toxic level and toxic responses were detected only by bioassays of amphipods and algae, but not by the Microtox1 test. The low toxicity is consistent with the low soluble content of contaminants of pore water and the low bioavailability of contaminants in sediments. # 1999 Elsevier Science Ltd. All rights reserved Key wordsÐpore water, sediment±water interface, contaminant bioavailability, toxicity assay, amphipods, algae, Microtox1 test, coastal wetland, the Mai Po Marshes
INTRODUCTION
The Mai Po Marshes Nature Reserve (22830'N and 114802'E) is situated in the north-western corner of the New Territories, Hong Kong, P.R. China, on the eastern edge of Deep Bay which borders the Shenzhen Special Economic Zone, Guangdong province to the north (Fig. 1). The Mai Po Marshes, which occupy an area of 381 hectares, together with the extensive intertidal mud¯ats and freshwater ®sh ponds, make up the largest remaining wetland in Hong Kong. In recent years, the Marshes have been under increasing pressure from environmental pollution and human disturbance. On the seaward side, the water quality of the Deep Bay is deteriorating, due to the increasing pollution loads from the Pearl River, the Shenzhen River, the Yuen Long Creek and the Kam Tin River. Economic development on both sides of the Shenzhen River has resulted in an increased discharge of industrial pollutants, which have progressively degraded the water quality of the Deep Bay. The Deep Bay, especially the inner subzone, *Author to whom all correspondence should be addressed. [Present address: Department of Geography, University of Cambridge, Downing Place, Cambridge, CB2 3EN, United Kingdom; Tel.: +44-1223-333375; Fax: +44-1223-333392; E-mail: [email protected]]. 909
has been claimed as one of the most polluted water bodies around Hong Kong, in terms of bacteria and nutrient loads, due to the discharges of domestic sewage, agricultural wastes derived from livestock rearing and industrial euents into the catchment areas (Hong Kong Environmental Protection Department, 1995a). Of these, the largest source of pollution is untreated domestic sewage and livestock wastewater, discharged via the Shenzhen River (Anonymous, 1994). Water quality in the Deep Bay is so poor that it not only exceeded the assimilative capacity of the water body (Hong Kong Environmental Protection Department, 1995b) but has also contributed to the decrease in shrimp productivity of the intertidal shrimp ponds (gei wais) in the wetland Mai Po Marshes in recent years (Cha et al., 1997). With the in¯ux of polluted water from the Deep Bay, contaminants are adsorbed onto suspended particles and subsequently accumulated in the underlying sediments in the Marshes (Lau, 1997). Because of the continual elevation of sediment level, the gei wais are dredged about every ®ve years to deepen the water channels in the pond to facilitate shrimp growth. Dredging will inevitably elevate the levels of suspended sediment in areas immediately adjacent to the dredging and may also lead to the potential release of contaminants, especially those loosely-bound, to the
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S. S. S. Lau and L. M. Chu
Fig. 1. Map of the Mai Po Marshes Nature Reserve, Hong Kong.
overlying water from the seriously contaminated sediments (Khalid et al., 1981; Gambrell, 1994; Tack et al., 1996). This threatens the biota and aects shrimp productivity in the gei wais. Most importantly, the gei wais are drained and dried for a week or two every year for the disinfection of the bottom sediment. This drying process may increase the mobility of contaminants released from the sediment to the overlying water when the gei wai is re¯ooded, thus augmenting the potential biological impacts on the ecosystem (Schoenberg and Oliver, 1988; Qiu and McComb, 1994). In the complex sediment±water system, the movement, availability and possible toxicity of contaminants are in¯uenced by chemical and physical factors like redox gradient, pH, salinity and temperature (Gambrell et al., 1991; Gambrell, 1994). Seasonal changes in these factors may therefore be important in governing the levels of bioavailable nutrients in the sediment±water interface. Amongst various physico±chemical properties of water, salinity and temperature are parameters which vary most widely with seasons in the estuarine gei wais (Lau, 1997). Knowledge on the mobilization of contaminants from the sediments in the shrimpgrowing gei wais under varying environmental con-
ditions (salinity and temperature) is essential in terms of ecological and conservational perspectives. The aims of this study were to (i) investigate the degree of contamination of sediments of Gei wai 13, (ii) examine the eects of various levels of salinity and temperature of the water on the release of contaminants from the sediment in the water± sediment system, (iii) examine the eect of drying on contaminant behavior of re¯ooded sediments and (iv) investigate the biological impacts of contaminants on the natural biota using bioassays of bacteria, algae and amphipods.
MATERIALS AND METHODS
Sediment collection The study was carried out in Gei wai 13, which is a shallow intertidal pond with vegetated platforms, at the Mai Po Marshes (Fig. 2). Triplicate undisturbed sediment samples were collected 5 m away from the sluice gate of the gei wai in March 1997 using a Van Veen grab (Fig. 3). Eight cm thick slices were taken and kept in plastic bags. Samples were kept at 48C in an ice box and analyzed within a week. The grab and plastic bags were acid washed, rinsed with distilled water and rinsed several times with seawater at the collection site before use.
Contaminant release from sediments in a coastal wetland
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Fig. 2. Location of Gei wai 13 at the Mai Po Marshes.
General physico-chemical analyses At the site, pH, redox potential (Eh) and the temperature of the sediment samples were measured directly, using a Checkmate watermeter (Corning, New York). The moisture content of the sediments was determined after the samples were oven-dried at 1058C for a week. Particle size distribution of the sediments was determined, using air-dried nonsieved samples by the hydrometer method. Total organic carbon (TOC) of sediments was determined by the Walkley±Black titration method. The concentrations of total nitrogen and total phosphorus were determined by a QickChem AE automated ion analyzer (Lachat Instruments, Milwaukee) after concentrated sulfuric acid digestion at 3608C. Total contents of heavy metals (cadmium, copper, nickel and zinc) were determined by a Hitachi Z-8100 atomic absorption spectrophotometer (Polarized Zeeman, Tokyo) after mixed acid (4:1 of conc. H2SO4: conc. HNO3) digestion. Attention was in particular given to pore-water of sediments, as pore-water in the sediment matrix may be forced out during compaction and consolidation and result in the release of contaminants into the water column. Pore-water of the sediment was obtained by centrifuging the sample at 10,000 rpm, at 48C, for 15 min and the ®ltered supernatant (pore-water) was collected for the measurement of salinity and the contents of nutrients and metals. Salinity was measured by a S-10 hand refractometer (Atago, Tokyo). Physico±chemical properties of the <63 mm (in diameter) sediment frac-
tion obtained by wet-sieving, using screen sieves were also determined. All analyses were performed according to standard methods (American Public Health Association, 1992). The pH, Eh, temperature and salinity of the overlying water were measured in situ, according to methods as described above. Eects of salinity and temperature The mobility of contaminants in sediments as a function of salinity, was studied at dierent temperatures in the laboratory. Triplicate wet undisturbed sediment samples were extracted with arti®cial seawater of dierent salinities (5, 15 and 25-) at a 1:1 (w/v) ratio at 15 and 258C using an orbital shaker (200 rpm, 1 h), followed by centrifugation at 10,000 rpm, 15 min. The levels of ammoniacal nitrogen, ortho-phosphate phosphorus, cadmium, copper, nickel and zinc in the extract were determined with suitable matrices according to standard methods (American Public Health Association, 1992). Eects of drying The eect of drying on contaminant mobility in sediments was studied by comparing intact wet sediment (as collected) with sediments, which had been air-dried at 208C in a ventilated incubator for 7 days. This is to simulate the drying of sediment in gei wais every year for the purpose of disinfection. The sediments were then extracted with arti®cial seawater of 15- salinity at 258C at a 1:1
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Fig. 3. Location of the sampling point at the Gei wai 13. (w/v) ratio by shaking at 200 rpm for 1 h. The treatments were replicated three times. The extraction condition of 25- and 258C was selected as it is close to the situation in the gei wai in November or December, during which they are drained and air-dried (Lau, 1997). The nutrient content (ammoniacal nitrogen and ortho-phosphate phosphorus) and metal content (cadmium, copper, nickel and zinc) in the extract were determined with suitable matrices according to standard methods (American Public Health Association, 1992). Toxicity assays Preparation of sediment extract for toxicity tests. Triplicate wet undisturbed sediment samples were extracted with 35- arti®cial seawater at 258C, at a 1:1 (w/v) ratio using an orbital shaker at 200 rpm for 1 h, followed by centrifugation at 10,000 rpm for 10 min. The supernatant was ®lter-sterilized using a 0.45 mm membrane ®lter and used for toxicity tests, which commenced within 2 days following extraction. Microtox1 test. The standard Microtox1 test employing a marine bacterial species, Vibrio ®scheri (previously named Photobacterium phosphoreum) was conducted to evaluate the toxicity of sediment extracts by a screening test using a Microtox M500 toxicity analyzer (Microbics, Carlsbad) (Microbics Corporation, 1995). The reduction of light emission from the bacteria after 5 and 15 min exposure at 158C was recorded (Casarini et al., 1991). The reduction in luminescence is proportional to the toxicity of the sample (Microbics Corporation, 1995). The whole sample was tested in the screening test. Zinc sulphate and
double distilled water were used as positive and negative controls, respectively (Microbics Corporation, 1995). Algal bioassay. The unicellular green alga Chlorella pyrenoidosa CU-2, originally isolated from the estuarine surface water of Tolo Harbour, Hong Kong (Chan, 1991) was used. Stock cultures were maintained on slants containing 2% agar in Walne's Medium at 22228C (Wang, 1996). Cultures from the agar slant were inoculated into 200 ml Walne's Medium in a 1000 ml Erlenmeyer ¯ask. Ten ml of algal cells at the logphase were then inoculated into 190 ml Walne's Medium in a 1000 ml Erlenmeyer ¯ask. The algal culture was incubated at 222 28C, under cool white ¯uorescent light, at an incident intensity of 4000 lux, as measured with a LI-189 photometer (Licor, NB). A 14/10 h light±dark cycle was used. For the algal growth inhibition test, sediment extract was serially diluted with autoclaved algal culture medium (Table 1) to result in 0, 6.25, 12.5, 25, 50 and 100% (v/v) dilutions. Triplicate 125 ml Erlenmeyer ¯asks of each test concentration, which contained 40 ml culture solution, were inoculated with C. pyrenoidosa CU-2 at logphase to achieve an initial concentration of 2 106 cells mlÿ1. The ¯asks were incubated in an incubation chamber at 22228C, illuminated with cool white ¯uorescent light, at an intensity of 4000 lux, with a 14/10 h light±dark cycle, on an orbital shaker operating at 100 rpm. Algal growth was monitored every day for ®ve days by measuring the absorbance of the culture suspension at a wavelength of 690 nm using a MR5000 microplate spectrometer (Dynatech Laboratories, Chantilly). One hundred ml of the well-mixed culture suspension was taken and added into a well of a 96 well microplate for absorbance measurement.
Contaminant release from sediments in a coastal wetland Table 1. Algal culture medium for Chlorella pyrenoidsa CU2 (modi®ed from Wang, 1996) Chemicals
Weight (g)
Preparation of stock solutions Solution Aa ZnCl2 CoCl2 6H2O (NH4)6Mo7O24 4H2O CuSO4 5H2O
2.100 2.000 0.900 2.000
Solution Bb Na2EDTA H3BO3 NaNO3 NaH2PO4 MnCl2 4H2O FeCl3
4.500 3.360 10.000 1.538 0.036 0.078
Solution Cc Thiamin hydrochloride Cyanocobalamin Arti®cial seawater Synthetic sea salt Tris base
0.200 0.010
d
30.000 2.430
The algal culture medium is prepared by adding 1 ml of solution B and 0.1 ml of solution C in to 1000 ml of autoclaved arti®cial seawater and mixed thoroughly. a The chemicals were dissolved in 100 ml of double distilled water and added a few drops of concentrated HCl to get dissolution well and autoclaved for 30 min at 1258C. b The chemicals were dissolved in 100 ml of a mixture of 1:99 solution A: double distilled water and autoclaved for 30 min at 1258C. c The chemicals were dissolved in 100 ml of double distilled water and ®lter-sterilized with 0.22 mm membrane ®lter. d The chemicals were dissolved in 1000 ml of double distilled water and adjusted at pH 7.6 with concentrated HCl and autoclaved for 30 min at 1258C.
All glassware was autoclaved before use and aseptic technique was practiced throughout the experiment. Amphipod bioassay. A local marine amphipod species Elasmopus rapax, originally collected from the Tolo Harbour, Hong Kong in 1993, was maintained in aerated ®ltered seawater in a 1000 l ®berglass tank. The water surface was partially covered by ®bre cotton to keep the system cool and dark. The temperature range for the species is between 10 and 338C, with an optimal range of 23±288C (Wang, 1996). The detail of the developmental process of E. rapax is not known, but the species had been used in toxicity tests before (Wang, 1996). It was observed that the amphipod can grow up to 12±14 mm in body length. The ®bre cotton to which the amphipods were attached, was gently washed with seawater in a plas-
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tic pail and the amphipods were then collected by passing the seawater through a 250 mm nylon net. Medium-sized (3±4 mm body length) amphipods were selected and kept in a 500 ml beaker with 300 ml seawater for two days without food supply, before being used for a toxicity test. Ten randomly selected amphipods were transferred to 300 ml plastic cups with a serial dilution of sediment extracts at 0, 6.25, 12.5, 25, 50 and 100% (v/v) dilutions. The sediment extract was diluted with arti®cial seawater. The cups were kept in an incubator at 25218C with a 14/10 h light±dark cycle for 96 h. Each treatment was replicated three times. No food or aeration was provided during the experiment. Mortality was recorded at 48 and 96 h and amphipods were de®ned as dead if no response to physical prodding was observed. Results of toxicity assays are expressed as percentage mortality. Statistical analyses In the toxicity assays with C. pyrenoidosa CU-2, maximum speci®c growth rate (mmax) was determined from the speci®c growth rate (m) data of each culture, calculated by the following equation: m = ln(n2ÿn1)/(t2ÿt1) where n1 and n2 were absorbances of the cultures at the beginning (t1) and at the end (t2) of the selected time interval (United States Environmental Protection Agency, 1971). Amphipod mortality was evaluated by regression analysis, using the probit program (Finney, 1971, 1978) and the median lethal concentration LC50 with 95% con®dence interval was determined, using the computer software package, Toxicologist V1.0 (1990) (the Euro-Mediterranean Center in Marine Contamination Hazards, Australia). The LC50 can only be calculated from experimental data, where two or more concentrations show partial mortalities and with a signi®cant w2 value in the w2-test (Finney, 1978). The Student's t-test, analysis of variance (ANOVA) and Tukey's honestly signi®cant dierence (HSD)-test for calculation of least signi®cant dierence (LSD) at p = 0.05 were carried out, whenever appropriate. All data were analyzed by SPSS (statistical package for social science) for Windows release 6.0, using an IBM Pentium PC. RESULTS AND DISCUSSION
General properties Table 2 presents the general properties of the overlying water and the pore-water of the sediment from Gei wai 13. The salinity of the pore-water was greater than that of the overlying water. Pore-water represents an important phase in the transport of contaminants from the sediment to the overlying water; the low soluble contents of nutrients and
Table 2. Physico-chemical properties of overlying- and pore-water of sediments sampled from Gei wai 13 Overlying water
pH Temperature (8C) Eh (mV) Salinity (-) NHx±N (mg lÿ1) Ortho-P (mg lÿ1) Cd (mg lÿ1) Cu (mg lÿ1) Ni (mg lÿ1) Zn (mg lÿ1)
Pore-water
mean
range
mean
range
7.18 25.7 ÿ9 12 NT NT NT NT NT NT
7.16±7.20 25.6±25.7 ÿ10 to ÿ9 12±12 NT NT NT NT NT NT
8.20 NT NT 16 10.8 3.73 ND 0.04 ND ND
8.19±8.21 NT NT 16±16 9.16±12.5 3.44±3.97 ÿ 0.04±0.05 ÿ ÿ
Value represents the mean of three replicates. NT denotes not tested.ND denotes not detectable (Cd < 0.05 mg lÿ1; Ni < 0.1 mg lÿ1; Zn < 0.04 mg lÿ1).
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S. S. S. Lau and L. M. Chu Table 3. Physico-chemical properties of bulk and <63 mm fraction sediments sampled from Gei wai 13 Sediment
p value
bulk
pH Temperature (8C) Eh (mV) Moisture content (%) Total organic carbon (%) Total N (mg kgÿ1) Total P (mg kgÿ1) Total Cd (mg kgÿ1) Total Cu (mg kgÿ1) Total Ni (mg kgÿ1) Total Zn (mg kgÿ1)
<63 mm fraction
mean
range
mean
range
7.32 24.7 ÿ22 62.0 2.10 1023 490 0.61 33.5 21.5 134
7.21±7.38 24.6±24.7 ÿ24 to ÿ20 56.4±69.6 1.92±2.33 950±1080 470±502 0.61±0.61 32.4±34.0 20.0±23.2 125±141
NT NT NT NT 0.86 403 210 0.68 39.5 32.8 177
NT NT NT NT 0.78±0.96 370±440 198±226 0.61±0.76 38.6±40.6 31.2±34.4 175±183
NA NA NA NA <0.01 <0.01 <0.01 NS <0.05 <0.01 <0.05
Value represents the mean of three replicates. All results of the contents of nutrients and metals are shown on dry weight. NT denotes not tested. NA denotes not applicable. NS denotes no signi®cant dierence (p = 0.05) by Student's t-test.
metals of pore water re¯ect a low potential mobility or biological impact of the contaminants of the sediments (Shaw et al., 1990; Hong et al., 1995). This shows that contaminants (nutrients and heavy metals) have strong chemical associations with the clay fraction and the organic material in the sediments and thus do not readily enter into the porewater. The impact on the water quality from the expelled pore-water contamination is thus minimal. The sediments consisted predominantly of silts and clays (about 90%) which were formed by the settlement of ®ne particulates from the silt-laden water from the Pearl River. Analyses of bulk sediment and the <63 mm fraction demonstrated, that most of the nutrient content (nitrogen and phosphorus) tended to bind to the coarser particles, while most of the metals (copper, nickel and zinc) bind to the ®ner particles (silts and clays) on which greater numbers of reactive sites were located (Table 3) (Harbison, 1986). The <63 mm fraction accounted for 40% of the total organic carbon content. Finer particles and organic matter are known to have a higher cation exchange capacity (Wiese et al., 1995) inferring that the sediments have a great ability to adsorb and concentrate metals. The degree of contamination of sediments in Gei wai 13 was compared with those sediments outside the Gei wai at the bottom of the water channel (Lau, 1997). Total concentrations of copper, nickel and zinc in the sediments from the water channel were nearly twice as high as those in Gei wai 13 and the total contents of nitrogen, phosphorus and cadmium were two times higher than those in Gei wai 13. These results suggest that sediment in Gei wai 13 was less contaminated than that from the water channel. However, anthropogenic enrichment of the sediment in Gei wai 13 was clearly evident, when the pollution level was compared with the ``background level'' in a wetland in Sai Keng, Hong Kong. Total contents of nitrogen, phosphorus and
cadmium of the sediment in Gei wai 13 was about twice as high as those in Sai Keng, while total contents of copper, nickel and zinc were about 10 times as high as those in Sai Keng (Tam and Wong, 1995). The dierence in contamination levels of sediments between Gei wai 13 and the water channel, can be attributed to the management regime of the sluice gate, which is the only opening of the Gei wai for tidal water exchange. The sluice gate opens manually only at spring tide for a period of 4±5 days a month, to let water ¯ush in to the Gei wai from the Deep Bay. Thus, in terms of exposure, sediment in Gei wai 13 was shielded by limited exposure to the polluted water from the Deep Bay and thus to the settlement of contaminated clayey particulates. Eects of temperature and salinity Figure 4 shows the concentrations of ammoniacal nitrogen and ortho-phosphate phosphorus of sediment extracts, under various levels of salinity and temperature. An increase in salinity from 5 to 15resulted in a marked increase (p < 0.05) in the levels of ammoniacal nitrogen of the sediment extracts at 258C, but there was no further increase when the salinity was raised to 25-. Changes in salinity at a lower temperature, 158C, had no signi®cant eect on the ammoniacal nitrogen level. This indicates, that the mobility of ammoniacal nitrogen of the sediment is increased, in response to an increase in salinity only under warm conditions (258C) (Pagnotta et al., 1989; Health, 1992). Signi®cant increase (p < 0.05) in the levels of ammoniacal nitrogen were found at salinities of 15 and 25- when the temperature was raised from 15 to 258C. A similar eect of salinity and temperature on the mobility of ortho-phosphate phosphorus of the sediment was noted. However signi®cant dierences in the levels of ortho-phosphate phosphorus
Contaminant release from sediments in a coastal wetland
915
Fig. 4. Concentrations of ammoniacal nitrogen (mg kgÿ1) and orthophosphate phosphorus (mg kgÿ1) in sediment extracts under various levels of salinity and temperature. Values are the means and standard deviations of triplicate. Vertical bar denotes LSD by Tukey's HSD test at p = 0.05.
between 15 and 258C were observed, only at a salinity of 25-, with the highest mobility of orthophosphate phosphorus being at 25- and 258C. An increase in temperature would enhance the desorption and dissolution ability of phosphorus, by the increase in the microbial activity in sediments (Bostrom et al., 1988). A statistically signi®cant temperature±salinity interaction was observed for ammoniacal nitrogen (F value = 8.02, p < 0.01) and ortho-phosphate phosphorus (F value = 3.89, p < 0.05) indicating a coupling eect of temperature and salinity on nutrient remobilization across the sediment±water interface. No comparison on the release of cadmium, copper, nickel and zinc from sediments could be made, since they were below the limits of the detection at all levels of salinity and temperature.
Eects of drying The water solubility of nutrients (nitrogen and phosphorus) in air-dried sediments was lower than those in wet sediments (Table 4). During the period of drying, the reduction in ammoniacal nitrogen may be accounted for by the nitri®cation of ammonium to nitrate (Groman and Tiedje, 1989; Yamaguchi et al., 1990). This sounds contradictory, as in some cases, drying treatment may lead to a release of ammoniacal nitrogen, either because of the microorganisms in the sediment being killed by desiccation (Sparling et al., 1985) or the organically-bound nitrogen being destroyed by oxidation (Kadlec and Knight, 1996). The reduction in the release of phosphorus from sediments, is possibly the result of the incorporation of phosphorus into
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S. S. S. Lau and L. M. Chu
Table 4. Water soluble portions of contaminants in wet and air-dried sediments sampled from Gei wai 13 (all values in mg kgÿ1 dry weight) Sediment
p value
wet
NHx±N Ortho-P Cd Cu Ni Zn
air-dried
mean
range
mean
range
16.2 2.12 ND 0.03 ND ND
14.2±17.7 2.06±2.20 ÿ 0.03±0.04 ÿ ÿ
9.30 0.60 ND 0.07 ND 0.05
8.48±9.86 0.57±0.64 ÿ 0.06±0.08 ÿ 0.04±0.05
<0.05 <0.01 NA <0.05 NA NA
Value represents the mean of three replicates. ND denotes not detectable. NA denotes not applicable.
aggregated particles during dehydration (Twinch, 1987) and the condensation of organic matter molecules, which results in greater accessibility of phosphate to adsorption sites on mineral colloids, that were previously obstructed by the presence of organic matter (Haynes and Swift, 1989). On the other hand, studies have shown that drying may lead to an increase in phosphorus release to the overlying water from the sediment (Bartlett and James, 1980; Schoenberg and Oliver, 1988; Qiu and McComb, 1994) by killing the microorganisms in sediments by desiccation (Sparling et al., 1985). The con¯icting literature of the potential drying eect on nutrient release, indicates the degree of complexity of the underlying mechanisms. In terms of metal release, the higher metal release (i.e. Cu and Zn) from the extracts of the air-dried sediment, compared with that of the wet sediment, is probably attributable to partial oxidation of the sediment organic matter during air-drying (Bartlett and James, 1980).
Toxicity assays Results from the toxicity assays are presented in Table 5. The Microtox1 test with V. ®scheri did
not indicate toxicity at 100% sediment extract, but stimulated bacterial growth, resulting in a higher light emission. It is likely that the presence of nutrients and trace metals allows for increased metabolism and thus light production in V. ®scheri, relative to responses in organisms maintained in the control medium. Amphipod survival at 48 h exhibited toxic responses to the sediment extracts, but no LC50 value could be obtained from the probit analysis. The 96 h LC50 value from the acute toxicity test with amphipods was 80.3% (48.7±234%), inferring a low toxic eect of the sediment extract. For C. pyrenoidosa CU-2, inhibitory eects of the sediment extracts on growth were shown in that a signi®cant dierence (p < 0.01) of maximum growth rate was observed in sediment extracts, with concentrations of r12.5%, except that there was an unexpectedly stimulatory response at 25% sediment extract. Results from the toxicity assays indicated, that the toxic substances in the sediment extracts appear to have greater impact on the survival of the amphipod and the algae than on the light production (i.e. metabolism) of the luminescent bacterium. It has to be recognized, that both the amphipod and the algae used in the toxicity testing, may have developed certain tolerance of polluted
Table 5. The results of acute toxicity tests of Microtox1, green algae and amphipod with sediment extract Sediment extract concentration (%)
Bioassays 1
0.0 6.25 12.5 25.0 50.0 100.0
Microtox (Vibrio ®scheri)
Green alga (Chlorella pyrenoidosa CU-2)
Amphipod (Elasmopus rapax)
luminescenceb
max. growth rate (dayÿ1)
mortality (%)
5 min
15 min
95 NT NT NT NT 175
95 NT NT NT NT 175
0.39 0.40 0.32a 0.47a 0.27a 0.14a
Value represents the mean of three replicates. NT denotes not tested. a indicates signi®cant dierence to control (0% sediment extract concentration) at p = 0.05 by Tukey's HSD test. b Luminescence measured at 0 min was 95.
48 h
96 h
0 0 3 7 17 37
0 10 20 33 43 53
Contaminant release from sediments in a coastal wetland
conditions, as they came from the grossly polluted Tolo Harbour. Moreover, the toxicity level in the tests, as indicated by the sediment extraction procedure, does not represent the only route of toxic exposure for aquatic organisms. Ingestion of contaminated sediment particulates, by infaunal organisms for instance, can be a major route of toxic exposure. Hence, low toxicity in the present study may not be the evidence of an absence of toxicity in the sediments. Furthermore, the eects of bioturbation (i.e. activities of ®sh or benthic invertebrates) on the sediment resuspension (e.g. Andersen and Jensen, 1991) in the shrimp ponds, should not be neglected in respect of remobilization and toxicity of the contaminants. CONCLUSIONS
In terms of the contents of nutrients and heavy metals, the sediment in Gei wai 13 was more contaminated than that of a remote wetland in Sai Keng. The sediment in Gei wai 13, however, was less contaminated, than the sediment in the water channel outside the gei wai in the Mai Po Marshes. With reference to the binding characteristics of the contaminants, nutrients (nitrogen and phosphorus) are likely to bind to the coarser particles while metals (copper, nickel and zinc) bind to the ®ner particles (silts and clays). Sediments with higher percentages of silts and clays had high concentrations of heavy metals. Seasonal change in Mai Po might play an important role aecting the mobility of contaminants from the sediment to the overlying water, in that nutrients in sediments were more liable to mobilization in saline water and in the hot season. This shows that following the reduction of external pollution loading by closing-up the sluice gates of gei wais, the water quality in the ponds could be impacted by the bottom nutrient-rich sediments. The enhanced nutrient (nitrogen and phosphorus) availability to the water, following sediment drying, implies the importance of management strategy after the annual pond drying. In contrast, the release of metals (i.e. Cu and Zn) to the extracts appeared to be greater in air-dried sediment than in wet sediment. Metal levels, measured by the ¯ame atomic absorption spectrophotometer, however, were close to or below the detection limits, further studies using a graphite furnace atomic absorption spectrophotometer for accurate metal measurement are recommended. Hence, understanding the underlying mechanisms responsible for the movement of contaminants across the sediment±water interface is important for the protection, conservation and management of the ecologically important wetlands. With high levels of contaminants however, sediment extract toxicity bioassays show low toxic impacts of the sediment, indicating a weak association between the content of contaminants and
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their potential toxicity. The low toxic response is consistent with the low soluble content of contaminants in the pore-water and the low availability of contaminants in sediments. In the present study, amphipods and algae were found to be more sensitive to the contaminated sediment than bacteria, as toxic responses were only detected in toxicity tests of the former two. The discrepancy between various toxicity tests indicates the necessity and importance of a multispecies approach in toxicological assessment. AcknowledgementsÐThanks are given to Dr P. K. Wong and Dr C. K. Wong of The Chinese University of Hong Kong for their kind donation of algae and amphipods for toxicity assays, respectively and Mr H. C. Ho, Miss P. K. Lee, Mr Y. K. Chan, Mr M. W. Tse and Dr L. Young for their technical assistance. Thanks are also due to Mr J. C. Green for proof-reading the manuscript. The manuscript bene®ted from criticism by three anomymous reviewers. The ®nancial support from the Chinese University of Hong Kong is gratefully acknowledged. REFERENCES
American Public Health Association (1992) Standard Methods for the Examination of Water and Wastewaters, 18th edn. American Public Health Association, Washington, DC. Andersen F. O. and Jensen H. S. (1991) The in¯uence of chironomids on decomposition of organic matter and nutrient exchange in a lake sediment. Verh. int. Ver. Limnol. 24, 3051±3055. Anonymous (1994) Environmental Impact Assessment Study on Shenzhen River Regulation Project: Preliminary Study Report. Beijing University, Beijing. Bartlett R. and James B. (1980) Studying dried, stored soil samples: some pitfalls. Soil Sci. Soc. Am. J. 44, 721± 724. Bostrom B., Anderson J. M., Fleischer S. and Jansson M. (1988) Exchange of phosphorus across the sediment± water interface. Hydrobiologia 170, 229±244. Casarini D. C. P., Cunha R. C. A., Sato M. I. Z. and Sanchez P. S. (1991) Evaluation of toxicity test procedure to de®ne loading rates in a land treatment system. Wat. Sci. Tech. 24, 183±188. Cha M. W., Young L. and Wong K. M. (1997) The fate of traditional extensive (gei wai) shrimp farming at the Mai Po Marshes Nature Reserve, Hong Kong. Hydrobiologia 352, 295±303. Chan L. C. (1991) Hydrography and marine plankton of Tolo Harbour, Hong Kong. M.Ph. thesis, The Chinese University of Hong Kong, Hong Kong. Finney D. J. (1971) Probit Analysis, 3rd edn. Cambridge University Press, London. Finney D. J. (1978) Statistical Method in Biological Assay. Charles Grin Company Limited, London. Gambrell R. P. (1994) Trace and toxic metals in wetlands: a review. J. Environ. Qual. 23, 883±891. Gambrell R. P., Wiesepape J. B., Patrick W. H. and Du M. C. (1991) The eects of pH, redox and salinity on metal release from a contaminated sediment. Water Air Soil Pollut. 57, 359±367. Groman P. M. and Tiedje J. M. (1989) Denitri®cation in north temperate forest soils: spatial and temporal patterns at the landscape level and seasonal scales. Soil Biol. Biochem. 21, 613±620. Harbison P. (1986) Mangrove muds: a sink and a source for trace metals. Mar. Pollut. Bull. 17, 246±250.
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Haynes R. J. and Swift R. S. (1989) The eects of pH and drying on adsorption of phosphate by aluminium±organic matter association. J. Soil Sci. 40, 773±781. Health R. T. (1992) Nutrient dynamics in Great Lakes coastal wetlands: future directions. J. Great Lakes Res. 18, 590±602. Hong J., Calmano W. and Forstner U. (1995) Interstitial waters. In Trace Elements in Natural Waters, ed. B. Salbu and E. Steinnes, pp. 117±176. CRC Press, Boca Raton. Hong Kong Environmental Protection Department (1995a) Environment Hong Kong 1995: A Review of 1994. Government Printer, Hong Kong. Hong Kong Environmental Protection Department (1995b) Marine Water Quality in Hong Kong for 1994. Government Printer, Hong Kong. Kadlec R. H. and Knight R. L. (1996) Wetland soils. In Treatment Wetlands, pp. 63±80. Lewis Publishers, Boca Raton. Khalid R. A., Gambrell R. P. and Patrick W. H., Jr. (1981) Chemical availability of cadmium in Mississippi River sediment. J. Environ. Qual. 10, 523±528. Lau S. S. S. (1997) Pollution status and assimilative potential of the wetlands at the Mai Po Marshes Nature Reserve, Hong Kong. M.Ph. thesis, The Chinese University of Hong Kong, Hong Kong. Microbics Corporation (1995) Microtox Acute Toxicity Screening Test Procedures. Microbics Corporation, Carlsbad. Pagnotta R., Blundo C. M., LaNoce T., Pettine M. and Puddu A. (1989) Nutrient remobilization processes at the Tiber river mouth (Italy). Hydrobiologia 176/177, 297±306. Qiu S. and McComb A. J. (1994) Eects of oxygen concentration on phosphorus release from re¯ooded airdried wetland sediments. Aust. J. Mar. Fresh. Res. 45, 1319±1328. Schoenberg S. A. and Oliver J. D. (1988) Temporal
dynamics and spatial variation of algae in relation to hydrology and sediment characteristics in the Okefenokee swamp, Georgia. Hydrobiologia 162, 123± 133. Shaw T. J., Gieskes J. M. and Jahnke R. A. (1990) Early diagnosis in diering depositional environments: the response of transition metals in pore water. Geochim. Cosmochim. Acta 54, 1233±1246. Sparling G. P., Whale K. N. and Ramsay A. J. (1985) Quantifying the contribution from the soil microbial biomass to the extractable P levels of fresh and air-dried soils. Aust. J. Soil Res. 23, 613±621. Tack F. M., Callewaert Q. W. J., Jr. and Verloo M. G. (1996) Metal solubility as function of pH in a contaminated, dredged sediment aected by oxidation. Environ. Pollut. 91, 199±208. Tam N. F. Y. and Wong Y. S. (1995) Mangrove soils as sinks for wastewater-borne pollutants. Hydrobiologia 295, 231±241. Twinch A. J. (1987) Phosphate exchange characteristics of wet and dried sediment from a hypertrophic reservoir: implications for the measurements of sediment phosphorus status. Water Res. 21, 1225±1230. United States Environmental Protection Agency (1971) Algal Assay Procedure, Bottle Test. USEPA, Corvallis. Wang J. (1996) Identi®cation and toxicological evaluation of polycyclic aromatic hydrocarbons in used crankcase oil. Ph.D. thesis, The Chinese University of Hong Kong, Hong Kong. Wiese S. B. O., Bubb J. M. and Lester J. N. (1995) The signi®cance of sediment metal concentrations in two eroding Essex salt marshes. Mar. Pollut. Bull. 30, 190±199. Yamaguchi T., Moldrup P., Teranishi S. and Rolston D. E. (1990) Denitri®cation in porous media during rapid, continuous leaching of synthetic wastewater at saturated water ¯ow. J. Environ. Qual. 19, 676± 683.
Wat. Res. Vol. 33, No. 4, pp. 909±918, 1999 # 1999 Elsevier Science Ltd. All rights reserved Printed in Great Britain 0043-1354/99/$ - see front matter
CONTAMINANT RELEASE FROM SEDIMENTS IN A COASTAL WETLAND M S. S. S. LAU* and L. M. CHU*
Department of Biology, The Chinese University of Hong Kong, Shatin, Hong Kong, P.R. China (First received February 1998; accepted in revised form July 1998) AbstractÐThe Mai Po Marshes Nature Reserve (Hong Kong) which is one of the most important regional wetland habitats, is under increasing pollution pressure, such that the existence and survival of the natural biota, including many endangered species, is threatened. Water quality in the tidal shrimp ponds (locally called gei wais) of the marshes was so poor, that biological impacts arising from the release of contaminants from the underlying sediments were suggested. A study was conducted to (i) investigate the level of contamination of underlying sediments in Gei wai 13, (ii) examine the eects of environmental factors on the behavior of contaminants in the sediments and (iii) study the toxicity of the sediments. Sediment in Gei wai 13 was contaminated with nutrients and heavy metals, of which nitrogen and phosphorus tend to bind to coarser particles, while copper, nickel and zinc to ®ner ones (silts and clays). Mobility of nitrogen and phosphorus from the sediment increased with increasing levels of salinity and temperature of the overlying water. In respect of the drying eect, the solubility of nitrogen and phosphorus of air-dried sediments was lower than that of wet sediments. Sediment extract toxicity tests indicate, that the highly contaminated sediment resulted in a low toxic level and toxic responses were detected only by bioassays of amphipods and algae, but not by the Microtox1 test. The low toxicity is consistent with the low soluble content of contaminants of pore water and the low bioavailability of contaminants in sediments. # 1999 Elsevier Science Ltd. All rights reserved Key wordsÐpore water, sediment±water interface, contaminant bioavailability, toxicity assay, amphipods, algae, Microtox1 test, coastal wetland, the Mai Po Marshes
INTRODUCTION
The Mai Po Marshes Nature Reserve (22830'N and 114802'E) is situated in the north-western corner of the New Territories, Hong Kong, P.R. China, on the eastern edge of Deep Bay which borders the Shenzhen Special Economic Zone, Guangdong province to the north (Fig. 1). The Mai Po Marshes, which occupy an area of 381 hectares, together with the extensive intertidal mud¯ats and freshwater ®sh ponds, make up the largest remaining wetland in Hong Kong. In recent years, the Marshes have been under increasing pressure from environmental pollution and human disturbance. On the seaward side, the water quality of the Deep Bay is deteriorating, due to the increasing pollution loads from the Pearl River, the Shenzhen River, the Yuen Long Creek and the Kam Tin River. Economic development on both sides of the Shenzhen River has resulted in an increased discharge of industrial pollutants, which have progressively degraded the water quality of the Deep Bay. The Deep Bay, especially the inner subzone, *Author to whom all correspondence should be addressed. [Present address: Department of Geography, University of Cambridge, Downing Place, Cambridge, CB2 3EN, United Kingdom; Tel.: +44-1223-333375; Fax: +44-1223-333392; E-mail: [email protected]]. 909
has been claimed as one of the most polluted water bodies around Hong Kong, in terms of bacteria and nutrient loads, due to the discharges of domestic sewage, agricultural wastes derived from livestock rearing and industrial euents into the catchment areas (Hong Kong Environmental Protection Department, 1995a). Of these, the largest source of pollution is untreated domestic sewage and livestock wastewater, discharged via the Shenzhen River (Anonymous, 1994). Water quality in the Deep Bay is so poor that it not only exceeded the assimilative capacity of the water body (Hong Kong Environmental Protection Department, 1995b) but has also contributed to the decrease in shrimp productivity of the intertidal shrimp ponds (gei wais) in the wetland Mai Po Marshes in recent years (Cha et al., 1997). With the in¯ux of polluted water from the Deep Bay, contaminants are adsorbed onto suspended particles and subsequently accumulated in the underlying sediments in the Marshes (Lau, 1997). Because of the continual elevation of sediment level, the gei wais are dredged about every ®ve years to deepen the water channels in the pond to facilitate shrimp growth. Dredging will inevitably elevate the levels of suspended sediment in areas immediately adjacent to the dredging and may also lead to the potential release of contaminants, especially those loosely-bound, to the
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Fig. 1. Map of the Mai Po Marshes Nature Reserve, Hong Kong.
overlying water from the seriously contaminated sediments (Khalid et al., 1981; Gambrell, 1994; Tack et al., 1996). This threatens the biota and aects shrimp productivity in the gei wais. Most importantly, the gei wais are drained and dried for a week or two every year for the disinfection of the bottom sediment. This drying process may increase the mobility of contaminants released from the sediment to the overlying water when the gei wai is re¯ooded, thus augmenting the potential biological impacts on the ecosystem (Schoenberg and Oliver, 1988; Qiu and McComb, 1994). In the complex sediment±water system, the movement, availability and possible toxicity of contaminants are in¯uenced by chemical and physical factors like redox gradient, pH, salinity and temperature (Gambrell et al., 1991; Gambrell, 1994). Seasonal changes in these factors may therefore be important in governing the levels of bioavailable nutrients in the sediment±water interface. Amongst various physico±chemical properties of water, salinity and temperature are parameters which vary most widely with seasons in the estuarine gei wais (Lau, 1997). Knowledge on the mobilization of contaminants from the sediments in the shrimpgrowing gei wais under varying environmental con-
ditions (salinity and temperature) is essential in terms of ecological and conservational perspectives. The aims of this study were to (i) investigate the degree of contamination of sediments of Gei wai 13, (ii) examine the eects of various levels of salinity and temperature of the water on the release of contaminants from the sediment in the water± sediment system, (iii) examine the eect of drying on contaminant behavior of re¯ooded sediments and (iv) investigate the biological impacts of contaminants on the natural biota using bioassays of bacteria, algae and amphipods.
MATERIALS AND METHODS
Sediment collection The study was carried out in Gei wai 13, which is a shallow intertidal pond with vegetated platforms, at the Mai Po Marshes (Fig. 2). Triplicate undisturbed sediment samples were collected 5 m away from the sluice gate of the gei wai in March 1997 using a Van Veen grab (Fig. 3). Eight cm thick slices were taken and kept in plastic bags. Samples were kept at 48C in an ice box and analyzed within a week. The grab and plastic bags were acid washed, rinsed with distilled water and rinsed several times with seawater at the collection site before use.
Contaminant release from sediments in a coastal wetland
911
Fig. 2. Location of Gei wai 13 at the Mai Po Marshes.
General physico-chemical analyses At the site, pH, redox potential (Eh) and the temperature of the sediment samples were measured directly, using a Checkmate watermeter (Corning, New York). The moisture content of the sediments was determined after the samples were oven-dried at 1058C for a week. Particle size distribution of the sediments was determined, using air-dried nonsieved samples by the hydrometer method. Total organic carbon (TOC) of sediments was determined by the Walkley±Black titration method. The concentrations of total nitrogen and total phosphorus were determined by a QickChem AE automated ion analyzer (Lachat Instruments, Milwaukee) after concentrated sulfuric acid digestion at 3608C. Total contents of heavy metals (cadmium, copper, nickel and zinc) were determined by a Hitachi Z-8100 atomic absorption spectrophotometer (Polarized Zeeman, Tokyo) after mixed acid (4:1 of conc. H2SO4: conc. HNO3) digestion. Attention was in particular given to pore-water of sediments, as pore-water in the sediment matrix may be forced out during compaction and consolidation and result in the release of contaminants into the water column. Pore-water of the sediment was obtained by centrifuging the sample at 10,000 rpm, at 48C, for 15 min and the ®ltered supernatant (pore-water) was collected for the measurement of salinity and the contents of nutrients and metals. Salinity was measured by a S-10 hand refractometer (Atago, Tokyo). Physico±chemical properties of the <63 mm (in diameter) sediment frac-
tion obtained by wet-sieving, using screen sieves were also determined. All analyses were performed according to standard methods (American Public Health Association, 1992). The pH, Eh, temperature and salinity of the overlying water were measured in situ, according to methods as described above. Eects of salinity and temperature The mobility of contaminants in sediments as a function of salinity, was studied at dierent temperatures in the laboratory. Triplicate wet undisturbed sediment samples were extracted with arti®cial seawater of dierent salinities (5, 15 and 25-) at a 1:1 (w/v) ratio at 15 and 258C using an orbital shaker (200 rpm, 1 h), followed by centrifugation at 10,000 rpm, 15 min. The levels of ammoniacal nitrogen, ortho-phosphate phosphorus, cadmium, copper, nickel and zinc in the extract were determined with suitable matrices according to standard methods (American Public Health Association, 1992). Eects of drying The eect of drying on contaminant mobility in sediments was studied by comparing intact wet sediment (as collected) with sediments, which had been air-dried at 208C in a ventilated incubator for 7 days. This is to simulate the drying of sediment in gei wais every year for the purpose of disinfection. The sediments were then extracted with arti®cial seawater of 15- salinity at 258C at a 1:1
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Fig. 3. Location of the sampling point at the Gei wai 13. (w/v) ratio by shaking at 200 rpm for 1 h. The treatments were replicated three times. The extraction condition of 25- and 258C was selected as it is close to the situation in the gei wai in November or December, during which they are drained and air-dried (Lau, 1997). The nutrient content (ammoniacal nitrogen and ortho-phosphate phosphorus) and metal content (cadmium, copper, nickel and zinc) in the extract were determined with suitable matrices according to standard methods (American Public Health Association, 1992). Toxicity assays Preparation of sediment extract for toxicity tests. Triplicate wet undisturbed sediment samples were extracted with 35- arti®cial seawater at 258C, at a 1:1 (w/v) ratio using an orbital shaker at 200 rpm for 1 h, followed by centrifugation at 10,000 rpm for 10 min. The supernatant was ®lter-sterilized using a 0.45 mm membrane ®lter and used for toxicity tests, which commenced within 2 days following extraction. Microtox1 test. The standard Microtox1 test employing a marine bacterial species, Vibrio ®scheri (previously named Photobacterium phosphoreum) was conducted to evaluate the toxicity of sediment extracts by a screening test using a Microtox M500 toxicity analyzer (Microbics, Carlsbad) (Microbics Corporation, 1995). The reduction of light emission from the bacteria after 5 and 15 min exposure at 158C was recorded (Casarini et al., 1991). The reduction in luminescence is proportional to the toxicity of the sample (Microbics Corporation, 1995). The whole sample was tested in the screening test. Zinc sulphate and
double distilled water were used as positive and negative controls, respectively (Microbics Corporation, 1995). Algal bioassay. The unicellular green alga Chlorella pyrenoidosa CU-2, originally isolated from the estuarine surface water of Tolo Harbour, Hong Kong (Chan, 1991) was used. Stock cultures were maintained on slants containing 2% agar in Walne's Medium at 22228C (Wang, 1996). Cultures from the agar slant were inoculated into 200 ml Walne's Medium in a 1000 ml Erlenmeyer ¯ask. Ten ml of algal cells at the logphase were then inoculated into 190 ml Walne's Medium in a 1000 ml Erlenmeyer ¯ask. The algal culture was incubated at 222 28C, under cool white ¯uorescent light, at an incident intensity of 4000 lux, as measured with a LI-189 photometer (Licor, NB). A 14/10 h light±dark cycle was used. For the algal growth inhibition test, sediment extract was serially diluted with autoclaved algal culture medium (Table 1) to result in 0, 6.25, 12.5, 25, 50 and 100% (v/v) dilutions. Triplicate 125 ml Erlenmeyer ¯asks of each test concentration, which contained 40 ml culture solution, were inoculated with C. pyrenoidosa CU-2 at logphase to achieve an initial concentration of 2 106 cells mlÿ1. The ¯asks were incubated in an incubation chamber at 22228C, illuminated with cool white ¯uorescent light, at an intensity of 4000 lux, with a 14/10 h light±dark cycle, on an orbital shaker operating at 100 rpm. Algal growth was monitored every day for ®ve days by measuring the absorbance of the culture suspension at a wavelength of 690 nm using a MR5000 microplate spectrometer (Dynatech Laboratories, Chantilly). One hundred ml of the well-mixed culture suspension was taken and added into a well of a 96 well microplate for absorbance measurement.
Contaminant release from sediments in a coastal wetland Table 1. Algal culture medium for Chlorella pyrenoidsa CU2 (modi®ed from Wang, 1996) Chemicals
Weight (g)
Preparation of stock solutions Solution Aa ZnCl2 CoCl2 6H2O (NH4)6Mo7O24 4H2O CuSO4 5H2O
2.100 2.000 0.900 2.000
Solution Bb Na2EDTA H3BO3 NaNO3 NaH2PO4 MnCl2 4H2O FeCl3
4.500 3.360 10.000 1.538 0.036 0.078
Solution Cc Thiamin hydrochloride Cyanocobalamin Arti®cial seawater Synthetic sea salt Tris base
0.200 0.010
d
30.000 2.430
The algal culture medium is prepared by adding 1 ml of solution B and 0.1 ml of solution C in to 1000 ml of autoclaved arti®cial seawater and mixed thoroughly. a The chemicals were dissolved in 100 ml of double distilled water and added a few drops of concentrated HCl to get dissolution well and autoclaved for 30 min at 1258C. b The chemicals were dissolved in 100 ml of a mixture of 1:99 solution A: double distilled water and autoclaved for 30 min at 1258C. c The chemicals were dissolved in 100 ml of double distilled water and ®lter-sterilized with 0.22 mm membrane ®lter. d The chemicals were dissolved in 1000 ml of double distilled water and adjusted at pH 7.6 with concentrated HCl and autoclaved for 30 min at 1258C.
All glassware was autoclaved before use and aseptic technique was practiced throughout the experiment. Amphipod bioassay. A local marine amphipod species Elasmopus rapax, originally collected from the Tolo Harbour, Hong Kong in 1993, was maintained in aerated ®ltered seawater in a 1000 l ®berglass tank. The water surface was partially covered by ®bre cotton to keep the system cool and dark. The temperature range for the species is between 10 and 338C, with an optimal range of 23±288C (Wang, 1996). The detail of the developmental process of E. rapax is not known, but the species had been used in toxicity tests before (Wang, 1996). It was observed that the amphipod can grow up to 12±14 mm in body length. The ®bre cotton to which the amphipods were attached, was gently washed with seawater in a plas-
913
tic pail and the amphipods were then collected by passing the seawater through a 250 mm nylon net. Medium-sized (3±4 mm body length) amphipods were selected and kept in a 500 ml beaker with 300 ml seawater for two days without food supply, before being used for a toxicity test. Ten randomly selected amphipods were transferred to 300 ml plastic cups with a serial dilution of sediment extracts at 0, 6.25, 12.5, 25, 50 and 100% (v/v) dilutions. The sediment extract was diluted with arti®cial seawater. The cups were kept in an incubator at 25218C with a 14/10 h light±dark cycle for 96 h. Each treatment was replicated three times. No food or aeration was provided during the experiment. Mortality was recorded at 48 and 96 h and amphipods were de®ned as dead if no response to physical prodding was observed. Results of toxicity assays are expressed as percentage mortality. Statistical analyses In the toxicity assays with C. pyrenoidosa CU-2, maximum speci®c growth rate (mmax) was determined from the speci®c growth rate (m) data of each culture, calculated by the following equation: m = ln(n2ÿn1)/(t2ÿt1) where n1 and n2 were absorbances of the cultures at the beginning (t1) and at the end (t2) of the selected time interval (United States Environmental Protection Agency, 1971). Amphipod mortality was evaluated by regression analysis, using the probit program (Finney, 1971, 1978) and the median lethal concentration LC50 with 95% con®dence interval was determined, using the computer software package, Toxicologist V1.0 (1990) (the Euro-Mediterranean Center in Marine Contamination Hazards, Australia). The LC50 can only be calculated from experimental data, where two or more concentrations show partial mortalities and with a signi®cant w2 value in the w2-test (Finney, 1978). The Student's t-test, analysis of variance (ANOVA) and Tukey's honestly signi®cant dierence (HSD)-test for calculation of least signi®cant dierence (LSD) at p = 0.05 were carried out, whenever appropriate. All data were analyzed by SPSS (statistical package for social science) for Windows release 6.0, using an IBM Pentium PC. RESULTS AND DISCUSSION
General properties Table 2 presents the general properties of the overlying water and the pore-water of the sediment from Gei wai 13. The salinity of the pore-water was greater than that of the overlying water. Pore-water represents an important phase in the transport of contaminants from the sediment to the overlying water; the low soluble contents of nutrients and
Table 2. Physico-chemical properties of overlying- and pore-water of sediments sampled from Gei wai 13 Overlying water
pH Temperature (8C) Eh (mV) Salinity (-) NHx±N (mg lÿ1) Ortho-P (mg lÿ1) Cd (mg lÿ1) Cu (mg lÿ1) Ni (mg lÿ1) Zn (mg lÿ1)
Pore-water
mean
range
mean
range
7.18 25.7 ÿ9 12 NT NT NT NT NT NT
7.16±7.20 25.6±25.7 ÿ10 to ÿ9 12±12 NT NT NT NT NT NT
8.20 NT NT 16 10.8 3.73 ND 0.04 ND ND
8.19±8.21 NT NT 16±16 9.16±12.5 3.44±3.97 ÿ 0.04±0.05 ÿ ÿ
Value represents the mean of three replicates. NT denotes not tested.ND denotes not detectable (Cd < 0.05 mg lÿ1; Ni < 0.1 mg lÿ1; Zn < 0.04 mg lÿ1).
914
S. S. S. Lau and L. M. Chu Table 3. Physico-chemical properties of bulk and <63 mm fraction sediments sampled from Gei wai 13 Sediment
p value
bulk
pH Temperature (8C) Eh (mV) Moisture content (%) Total organic carbon (%) Total N (mg kgÿ1) Total P (mg kgÿ1) Total Cd (mg kgÿ1) Total Cu (mg kgÿ1) Total Ni (mg kgÿ1) Total Zn (mg kgÿ1)
<63 mm fraction
mean
range
mean
range
7.32 24.7 ÿ22 62.0 2.10 1023 490 0.61 33.5 21.5 134
7.21±7.38 24.6±24.7 ÿ24 to ÿ20 56.4±69.6 1.92±2.33 950±1080 470±502 0.61±0.61 32.4±34.0 20.0±23.2 125±141
NT NT NT NT 0.86 403 210 0.68 39.5 32.8 177
NT NT NT NT 0.78±0.96 370±440 198±226 0.61±0.76 38.6±40.6 31.2±34.4 175±183
NA NA NA NA <0.01 <0.01 <0.01 NS <0.05 <0.01 <0.05
Value represents the mean of three replicates. All results of the contents of nutrients and metals are shown on dry weight. NT denotes not tested. NA denotes not applicable. NS denotes no signi®cant dierence (p = 0.05) by Student's t-test.
metals of pore water re¯ect a low potential mobility or biological impact of the contaminants of the sediments (Shaw et al., 1990; Hong et al., 1995). This shows that contaminants (nutrients and heavy metals) have strong chemical associations with the clay fraction and the organic material in the sediments and thus do not readily enter into the porewater. The impact on the water quality from the expelled pore-water contamination is thus minimal. The sediments consisted predominantly of silts and clays (about 90%) which were formed by the settlement of ®ne particulates from the silt-laden water from the Pearl River. Analyses of bulk sediment and the <63 mm fraction demonstrated, that most of the nutrient content (nitrogen and phosphorus) tended to bind to the coarser particles, while most of the metals (copper, nickel and zinc) bind to the ®ner particles (silts and clays) on which greater numbers of reactive sites were located (Table 3) (Harbison, 1986). The <63 mm fraction accounted for 40% of the total organic carbon content. Finer particles and organic matter are known to have a higher cation exchange capacity (Wiese et al., 1995) inferring that the sediments have a great ability to adsorb and concentrate metals. The degree of contamination of sediments in Gei wai 13 was compared with those sediments outside the Gei wai at the bottom of the water channel (Lau, 1997). Total concentrations of copper, nickel and zinc in the sediments from the water channel were nearly twice as high as those in Gei wai 13 and the total contents of nitrogen, phosphorus and cadmium were two times higher than those in Gei wai 13. These results suggest that sediment in Gei wai 13 was less contaminated than that from the water channel. However, anthropogenic enrichment of the sediment in Gei wai 13 was clearly evident, when the pollution level was compared with the ``background level'' in a wetland in Sai Keng, Hong Kong. Total contents of nitrogen, phosphorus and
cadmium of the sediment in Gei wai 13 was about twice as high as those in Sai Keng, while total contents of copper, nickel and zinc were about 10 times as high as those in Sai Keng (Tam and Wong, 1995). The dierence in contamination levels of sediments between Gei wai 13 and the water channel, can be attributed to the management regime of the sluice gate, which is the only opening of the Gei wai for tidal water exchange. The sluice gate opens manually only at spring tide for a period of 4±5 days a month, to let water ¯ush in to the Gei wai from the Deep Bay. Thus, in terms of exposure, sediment in Gei wai 13 was shielded by limited exposure to the polluted water from the Deep Bay and thus to the settlement of contaminated clayey particulates. Eects of temperature and salinity Figure 4 shows the concentrations of ammoniacal nitrogen and ortho-phosphate phosphorus of sediment extracts, under various levels of salinity and temperature. An increase in salinity from 5 to 15resulted in a marked increase (p < 0.05) in the levels of ammoniacal nitrogen of the sediment extracts at 258C, but there was no further increase when the salinity was raised to 25-. Changes in salinity at a lower temperature, 158C, had no signi®cant eect on the ammoniacal nitrogen level. This indicates, that the mobility of ammoniacal nitrogen of the sediment is increased, in response to an increase in salinity only under warm conditions (258C) (Pagnotta et al., 1989; Health, 1992). Signi®cant increase (p < 0.05) in the levels of ammoniacal nitrogen were found at salinities of 15 and 25- when the temperature was raised from 15 to 258C. A similar eect of salinity and temperature on the mobility of ortho-phosphate phosphorus of the sediment was noted. However signi®cant dierences in the levels of ortho-phosphate phosphorus
Contaminant release from sediments in a coastal wetland
915
Fig. 4. Concentrations of ammoniacal nitrogen (mg kgÿ1) and orthophosphate phosphorus (mg kgÿ1) in sediment extracts under various levels of salinity and temperature. Values are the means and standard deviations of triplicate. Vertical bar denotes LSD by Tukey's HSD test at p = 0.05.
between 15 and 258C were observed, only at a salinity of 25-, with the highest mobility of orthophosphate phosphorus being at 25- and 258C. An increase in temperature would enhance the desorption and dissolution ability of phosphorus, by the increase in the microbial activity in sediments (Bostrom et al., 1988). A statistically signi®cant temperature±salinity interaction was observed for ammoniacal nitrogen (F value = 8.02, p < 0.01) and ortho-phosphate phosphorus (F value = 3.89, p < 0.05) indicating a coupling eect of temperature and salinity on nutrient remobilization across the sediment±water interface. No comparison on the release of cadmium, copper, nickel and zinc from sediments could be made, since they were below the limits of the detection at all levels of salinity and temperature.
Eects of drying The water solubility of nutrients (nitrogen and phosphorus) in air-dried sediments was lower than those in wet sediments (Table 4). During the period of drying, the reduction in ammoniacal nitrogen may be accounted for by the nitri®cation of ammonium to nitrate (Groman and Tiedje, 1989; Yamaguchi et al., 1990). This sounds contradictory, as in some cases, drying treatment may lead to a release of ammoniacal nitrogen, either because of the microorganisms in the sediment being killed by desiccation (Sparling et al., 1985) or the organically-bound nitrogen being destroyed by oxidation (Kadlec and Knight, 1996). The reduction in the release of phosphorus from sediments, is possibly the result of the incorporation of phosphorus into
916
S. S. S. Lau and L. M. Chu
Table 4. Water soluble portions of contaminants in wet and air-dried sediments sampled from Gei wai 13 (all values in mg kgÿ1 dry weight) Sediment
p value
wet
NHx±N Ortho-P Cd Cu Ni Zn
air-dried
mean
range
mean
range
16.2 2.12 ND 0.03 ND ND
14.2±17.7 2.06±2.20 ÿ 0.03±0.04 ÿ ÿ
9.30 0.60 ND 0.07 ND 0.05
8.48±9.86 0.57±0.64 ÿ 0.06±0.08 ÿ 0.04±0.05
<0.05 <0.01 NA <0.05 NA NA
Value represents the mean of three replicates. ND denotes not detectable. NA denotes not applicable.
aggregated particles during dehydration (Twinch, 1987) and the condensation of organic matter molecules, which results in greater accessibility of phosphate to adsorption sites on mineral colloids, that were previously obstructed by the presence of organic matter (Haynes and Swift, 1989). On the other hand, studies have shown that drying may lead to an increase in phosphorus release to the overlying water from the sediment (Bartlett and James, 1980; Schoenberg and Oliver, 1988; Qiu and McComb, 1994) by killing the microorganisms in sediments by desiccation (Sparling et al., 1985). The con¯icting literature of the potential drying eect on nutrient release, indicates the degree of complexity of the underlying mechanisms. In terms of metal release, the higher metal release (i.e. Cu and Zn) from the extracts of the air-dried sediment, compared with that of the wet sediment, is probably attributable to partial oxidation of the sediment organic matter during air-drying (Bartlett and James, 1980).
Toxicity assays Results from the toxicity assays are presented in Table 5. The Microtox1 test with V. ®scheri did
not indicate toxicity at 100% sediment extract, but stimulated bacterial growth, resulting in a higher light emission. It is likely that the presence of nutrients and trace metals allows for increased metabolism and thus light production in V. ®scheri, relative to responses in organisms maintained in the control medium. Amphipod survival at 48 h exhibited toxic responses to the sediment extracts, but no LC50 value could be obtained from the probit analysis. The 96 h LC50 value from the acute toxicity test with amphipods was 80.3% (48.7±234%), inferring a low toxic eect of the sediment extract. For C. pyrenoidosa CU-2, inhibitory eects of the sediment extracts on growth were shown in that a signi®cant dierence (p < 0.01) of maximum growth rate was observed in sediment extracts, with concentrations of r12.5%, except that there was an unexpectedly stimulatory response at 25% sediment extract. Results from the toxicity assays indicated, that the toxic substances in the sediment extracts appear to have greater impact on the survival of the amphipod and the algae than on the light production (i.e. metabolism) of the luminescent bacterium. It has to be recognized, that both the amphipod and the algae used in the toxicity testing, may have developed certain tolerance of polluted
Table 5. The results of acute toxicity tests of Microtox1, green algae and amphipod with sediment extract Sediment extract concentration (%)
Bioassays 1
0.0 6.25 12.5 25.0 50.0 100.0
Microtox (Vibrio ®scheri)
Green alga (Chlorella pyrenoidosa CU-2)
Amphipod (Elasmopus rapax)
luminescenceb
max. growth rate (dayÿ1)
mortality (%)
5 min
15 min
95 NT NT NT NT 175
95 NT NT NT NT 175
0.39 0.40 0.32a 0.47a 0.27a 0.14a
Value represents the mean of three replicates. NT denotes not tested. a indicates signi®cant dierence to control (0% sediment extract concentration) at p = 0.05 by Tukey's HSD test. b Luminescence measured at 0 min was 95.
48 h
96 h
0 0 3 7 17 37
0 10 20 33 43 53
Contaminant release from sediments in a coastal wetland
conditions, as they came from the grossly polluted Tolo Harbour. Moreover, the toxicity level in the tests, as indicated by the sediment extraction procedure, does not represent the only route of toxic exposure for aquatic organisms. Ingestion of contaminated sediment particulates, by infaunal organisms for instance, can be a major route of toxic exposure. Hence, low toxicity in the present study may not be the evidence of an absence of toxicity in the sediments. Furthermore, the eects of bioturbation (i.e. activities of ®sh or benthic invertebrates) on the sediment resuspension (e.g. Andersen and Jensen, 1991) in the shrimp ponds, should not be neglected in respect of remobilization and toxicity of the contaminants. CONCLUSIONS
In terms of the contents of nutrients and heavy metals, the sediment in Gei wai 13 was more contaminated than that of a remote wetland in Sai Keng. The sediment in Gei wai 13, however, was less contaminated, than the sediment in the water channel outside the gei wai in the Mai Po Marshes. With reference to the binding characteristics of the contaminants, nutrients (nitrogen and phosphorus) are likely to bind to the coarser particles while metals (copper, nickel and zinc) bind to the ®ner particles (silts and clays). Sediments with higher percentages of silts and clays had high concentrations of heavy metals. Seasonal change in Mai Po might play an important role aecting the mobility of contaminants from the sediment to the overlying water, in that nutrients in sediments were more liable to mobilization in saline water and in the hot season. This shows that following the reduction of external pollution loading by closing-up the sluice gates of gei wais, the water quality in the ponds could be impacted by the bottom nutrient-rich sediments. The enhanced nutrient (nitrogen and phosphorus) availability to the water, following sediment drying, implies the importance of management strategy after the annual pond drying. In contrast, the release of metals (i.e. Cu and Zn) to the extracts appeared to be greater in air-dried sediment than in wet sediment. Metal levels, measured by the ¯ame atomic absorption spectrophotometer, however, were close to or below the detection limits, further studies using a graphite furnace atomic absorption spectrophotometer for accurate metal measurement are recommended. Hence, understanding the underlying mechanisms responsible for the movement of contaminants across the sediment±water interface is important for the protection, conservation and management of the ecologically important wetlands. With high levels of contaminants however, sediment extract toxicity bioassays show low toxic impacts of the sediment, indicating a weak association between the content of contaminants and
917
their potential toxicity. The low toxic response is consistent with the low soluble content of contaminants in the pore-water and the low availability of contaminants in sediments. In the present study, amphipods and algae were found to be more sensitive to the contaminated sediment than bacteria, as toxic responses were only detected in toxicity tests of the former two. The discrepancy between various toxicity tests indicates the necessity and importance of a multispecies approach in toxicological assessment. AcknowledgementsÐThanks are given to Dr P. K. Wong and Dr C. K. Wong of The Chinese University of Hong Kong for their kind donation of algae and amphipods for toxicity assays, respectively and Mr H. C. Ho, Miss P. K. Lee, Mr Y. K. Chan, Mr M. W. Tse and Dr L. Young for their technical assistance. Thanks are also due to Mr J. C. Green for proof-reading the manuscript. The manuscript bene®ted from criticism by three anomymous reviewers. The ®nancial support from the Chinese University of Hong Kong is gratefully acknowledged. REFERENCES
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