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Impact of heavy metals on denitrification in surface wetland sediments receiving wastewater
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Wal SCI Tech Vol. 40, No 3, pp. 349-355, I
Pergamon
Printed in Great Bntam. All nghts reset 0273-1223/99 $20.00 + I
PH: S0273-1223(99)00471-0
IMPACT OF REAVY METALS ON DENITRIFICATION IN SURFACE WETLAND SEDIMENTS RECEIVING WASTEWATER K. Sakadevan*, Huang Zheng** and H. J. Bavor* • Water Research Laboratory. University of Western Sydney Hawkesbury, Richmond, NSW-2753, Australia •• Institute ofEnvironmental Medicine, Medical University. Wuhan 430030, P.R. China
ABSTRACT Denitrification in sediment-water systems is a predominant process in the removal of nitrogen from wetlands and sediments receiving recycled water. In this study the impact ofcadmium (Cd), copper (Cu) and zinc (Zn) on denitrification was examined for a wetland sediment receiving recycled water. Results from the study showed that application of 100 mg Cd kg" sediment had no effect (P<0.05) on denitrification (161±2.7 mg N kg" sediment) compared to the control (162±2.4 mg N kg" sediment) which did not receive Cd, Cu or Zn. Addition of )00 mg Cu or Zn kg" sediment significantly increased denitrification (170± 1.8 and 168±2.7 mg NzO-N kg" sediment for Cu and Zn, respectively) compared to the control treatment. Addition of Cd, Cu or Zn at 500 or 1000 mg kg" sediment significantly decreased (P<0.05) total denitrification compared to the control and treatments. which received 100 mg Cd, Cu or Zn kg" sediment. For a given heavy metal concentration the largest denitrification inhibition occurred with Cd (30.9%) followed by Zn (24.9%) and Cu (18.9%) over a period of seven days. The amount of ammonium in the sediment water increased 1D all treatments receiving Cd, Cu or Zn and the concentration increased as the concentration of Cd, Cu or Zn increased in the sediment-water environment. For a given heavy metal concentration, the largest increase in ammonium occurred in treatments receiving Cd (31.1±0.9 mg N kg" sediment) followed by Zn (24.8±0.5 mg N kg" sediment) and Cu (l7.0±0.3 mg N kg" sediment). Denitrification inhibition was linearly related to the concentration of ammonium in sediment water (r=0.928). In general, the study showed that the addition of Cd, Cu or Zn inhibited denitrification and increased the concentration of ammonium in the sediment-water environment. iCI 1999 Published by Elsevier Science Ltd on behalf of the IAWQ. All nghts reserved
KEYWORDS Denitrification; heavy metal; inhibition; recycled water; sediment; run-off water; wetlands. INTRODUCTION Constructed wetland treatments of recycled and runoff waters are currently being practiced in man) countries as an alternative to the direct disposal of recycled or runoff water to coastal and inland waters (Bavor and Mitchell, 1994; Roser and Bavor, 1994; Sakadevan and Bavor, 1998). Recycled water and agricultural and urban community storm water runoff contain high concentrations of nitrogen (N) in the forms of oxidised nitrogen, ammonium and organic N. The excess concentration of N along with P ir recycled and run-off waters may cause eutrophication of surface waters leading to accelerated plant growth, changes in plant species composition and decrease the dissolved oxygen concentration of water (Zann 349
K. SAKADEVAN et al.
350
1995). Nitrate is also one of the most common contaminants in ground water and high concentrations of nitrate have been recognised as a health hazard in drinking water (Freeze and Cherry, 1979; Magee, 1982). When recycled or run-off water is applied to wetland systems, the ammonium and organic N are converted to nitrate through mineralisation and nitrification. Not being strongly adsorbed by sediment, nitrate may move to surface and ground waters. To protect ground and surface waters, the excess nitrate should be minimised via denitrification processes. The nitrification-denitrification process in sediment-water systems is considered to be a predominant process in the removal ofN from sediment-water systems (Korom, 1992). Denitrification in surface sediments receiving recycled water can prevent the movement of nitrate to ground and surface waters (Gilliam, 1994). In this biological process, ammonium is first oxidised to nitrate and then sequentially reduced to molecular N by denitrifying organisms. The process may be inhibited by sustained loads of various potentially toxic compounds including heavy metals, pesticide residues and other organic compounds such as polychlorinated biphenyls in recycled water, which are toxic to denitrifying organisms. An accumulation of heavy metals by long term application of biosolids may decrease the activity of sediment organisms (McGrath et al., 1988). Some effluent and biosolids contain metal concentrations up to 3,400 mg Cd, 10,400 mg Cu and 27,800 mg Zn kg'! dry weight biosolid (Miller and Donahue, 1990), which when applied to sediment may result in accumulation of large amounts of Cd, Cu and Zn in the sediment and may affect denitrification. The objective of the present study is to examine the effect of cadmium (Cd), copper (Cu) and zinc (Zn) and their concentrations on denitrification rates in wetland sediments and its relationship to N removal processes in the treatment of recycled water. MATERIALS AND METHODS Sediment The sediment used in the study was collected from a wetland drain on the University of Western Sydney Hawkesbury Dairy-Farm, situated 60 km northwest of Sydney, NSW, Australia. The farm is currently irrigated with recycled water obtained from the Richmond Sewage Treatment Plant, situated very close to the farm, and the wetland drain receives runoff from the irrigation. Sediment samples (50 cores of the 150 mm surface sediment, within a transect of 50 m x 10m) were collected from one wetland drain of the irrigated farm. All 50 cores of field moist sediment samples were bulked and sieved to less than 2 mm and stored at 4°C. Some chemical characteristics of the sediment are given in Table 1. Table 1. Chemical characteristics ofthe sediment Total C (g C kg,l sediment) Total N (g N kg,l sediment) Total P (mg P kg,lsediment) pH (l :2.5 H20) Extractable nitrate (mg N kg" sediment) Extractable ammonium ( mg kg" sediment) Exchangeable calcium (C mole(+) kg" sediment) Exchangeable magnesium (C mole(+~ kg') sediment) Exchangeable sodium (C mole(+)kg' sediment) Sodium Adsorption Ratio Total Cd (mg Cd kg" sediment) Total Cu (mg Cu kg" sediment) Total Zn (mg Zn kg"\ sediment)
16.6 1.4 603 6.3 4.8 3.1 2.55 1.83 0.91 0.62 1.2 4.8 17.8
Incubation experiment Field moist sediment samples (equivalent to 10 g dry weight) were weighed into 120 ml serum vials and the vials were treated separately with 1 ml of glucose solution (carbon source) containing 1800 g C L'I, 1 ml of heavy metal solution (separate heavy metal solutions for different treatments) and 7 ml deionised water. Including control, a total of 10 treatments were established and three replications were used for each treatment. The individual vials were then closed with rubber suba seal serum caps. A gentle stream of N2
Impact of heavy metals on denitrification
351
(300 ml minute") was directed over the slurry for three minutes by using an inlet and outlet needle, which pierced through the septum. After directing N2 over the slurry, 5 ml of head-space N2 from the vial was replaced with 5 ml of acetylene (C2H2) using a gas-tight syringe. Finally, 1 ml of sodium nitrate containing 1800 mg N L was injected to the vial to give a final concentration of 180 mg N Lo 1 in solution or 180 mg N kg" sediment. The final C concentration in solution also was 180 mg C LO\ or 180mg C kg" sediment. The control treatment was identical to other treatment except it received 1 ml of water instead of heavy metal solution. The vials were gently shaken for one hour and then incubated at room temperature. O
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Sediment, water and gaseous measurement The total C content of the sediment was measured before the experiment using a carbon-nitrogen-sulphur analyser (Fison, NA 1500 NCS). The total N content of the sediment was measured using an Alpkem Flow Solution 111 flow injection analyser (APHA, 1992) after digesting the sediment with a mixture of sulphuric acid, potassium sulphate and selenium (Kjeldahl mixture). The mineral N (nitrate + ammonium) in the sediment was measured using an Alpkem Flow Solution III system as above, after extracting the sediment with 2M KCl (1:10 ratio). After the experiment, the sediment water was filtered and the mineral N (nitrate + ammonium) concentration of the filtrate was measured as above. The sediment denitrification was estimated by the acetylene inhibition technique to yield a potential denitrification rate. At 24-hour intervals, a 0.2 ml headspace sample was removed from each vial using a gas-tight syringe and nitrous oxide (N20) emission from the sediment-water system was analysed by gas chromatography (GC) with a thermal conductivity detector. The denitrification inhibition rate was calculated as follows:
Statistical analysis Data were analysed using the analysis of variance (ANOVA) of SysStat using heavy metal type and concentration as the main effects on denitrification rate and ammonium and nitrate concentrations in sediment water at the end of the experiment. The interaction between three heavy metal concentrations on denitrification and ammonium formation was examined. Differences between the means within each effect were evaluated using a Fisher's protected least-significant-difference test. RESULTS AND DISCUSSION Impacts of Cd. Cu and Zn on denitrification The cumulative amount of N20 released with time for all treatments showed that significant (P
352
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Time (days) Figure I. Cumulative amount ofN20 oxide produced for sediment receiving (a) cadmium; (b) copper; and (c) zinc at no metal (-8-) and addition of 100 (--A-), 500 ( ) and 1,000 (__) mg metal kg" sediment.
The decrease in N20 released was greater (P
Impact of heavy metals on denitrification
353
than for Cu or Zn for a given concentration. Copper and Zn had similar effects on denitrification and N20 release from sediment-water which received 100 and 500 mg Cu or Zn kg'! sediment, but the decrease in N20 was greater for treatment which received 1000 mg Zn kg'! sediment than treatment that received 1000 mg Cu kg" sediment. Inhibitory effects were observed at 500 and 1000 mg Cd, Cu and Zn kg" sediment amendments (Fig. 2), and indicated that the order of inhibition was Cd>Zn>Cu for all levels of addition.
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Figure 2. Denitrification inhibition in sediment receiving 500 ( ) and 1000 (~) mg Cd kg" sediment, 500 (~) and 1000 (-e-) mg Cu kg" sediment and 500 (-9-) and 1000 (-+-) mg Zn kg' sediment.
The denitrification inhibition decreased during the incubation period (Fig. 2). This decrease may be due to: (1) among the wide range of denitrifying micro-organisms, at least a portion of the original population was able to overcome the inhibitory effects and developed tolerance to Cd, Cu or Zn. Similar phenomenon was observed that adaptation to heavy metals and selection of the resistant strains could occur rapidly in the sediments among a variety of nitrifying micro-organisms (Rother et al., 1982) and denitrifying population was also capable of recovering from the inhibitory effects ofthe toxicants (Bergeron et al., 1993); and/or (2) the amount of the bioavailable Cd, Cu or Zn decreased in the sediment-water due to adsorption, chelation and precipitation processes in the sediment (Ross, 1994). Ammonium and nitrate in sediment solution The concentration of ammonium in sediment solution in all treatments including the control was increased at the end of the experiment (Table 2), while nitrate concentrations decreased in all treatments (including control), even though about 180 mg nitrate N kg" sediment was added to the sediment at the beginning of the experiment (Table 2). The increases in ammonium concentrations were significantly greater in treatments which received 500 or 1000 mg Cd, Cu or Zn kg'! sediment compared to the control (P<0.05) and treatments which received 100 mg kg'! sediment of Cd, Cu or Zn. There was no significant difference in the amount of ammonium in sediment solution between control and treatments amended with 100 mg Cd, Cu or Zn kg" sediment (Table 2). The largest ammonium accumulation was observed in treatments which received 1000 mg Cd, Cu or Zn kg'! sediment. For a given concentration, the largest ammonium accumulation was observed for Cd followed by Zn and Cu (Table 2). With the nitrate reduction to N 2 0 and N 2 by denitrification processes, dissimilatory reduction of nitrate to ammonium N could have occurred in the same anoxic and carbon rich sediment-water environment, mediated by spore forming genera of Clostridium and Bacillus (Buresh and Patrick, 1978; Caskey and Tiedje, 1979). Both denitrification and the dissimilatory nitrate reduction are processes in which they make energy available to the cell for growth and maintenance (Harris, 1982). The denitrification inhibition by Cd, Cu or Zn may have resulted in some nitrate diversion into the pathway of reduction to ammonium rather tpan N20. This would result in more ammonium accumulation in sediment systems containing 500 or 1000 mg Cd, Cu or Zn kg'! sediment than control and treatments which received 100 mg Cd, Cu or Zn kg" sediment (Table 2). It was observed in the study that all added nitrate had not been denitrified (lower N 20 measured than nitrate added to the sediment-water environment) in sediment-water system of all treatments. Mass balance calculations of nitrate and
K. SAKADEVAN et al.
354
ammonium showed that the mineral N (sum of nitrate and ammonium) concentrations in sediment solutions at the end of the experiment had decreased for all treatments, reflecting a net loss of ammonium and/or nitrate from sediment solution (Table 2). This "loss" from sediment solution is hypothesised to have occurred through ammonia volatilisation or conversion to organic N by microbial biomass uptake. Table 2. Concentration of ammonium and nitrate in sediment water at the end of the experiment for all treatments Net mineral N lost Concentration Ammonium Nitrate Metal from sediment solution mg N kg" sediment
Cu
Zn
Control
7.8:1::0.85 17.1:1::2.43 31.1:1::0.72 6.1:1::0.70 11.5:1::1.08 17.0±0.32 5.8:1::0.90 15.1:1::1.42 24.8:1::0.52 9.6:1::0.82
100 500 1000 100 500 1000 100 500 1000
Cd
No added metal
1.1:1::0.15 1.5:1::0.36 1.5:1::0.06 0.9±0.4 1.2:1::0.06 0.8:1::0.06 1.0±.11 0.9±0.06 1.0±0.21 0.8:1::0.06
17.8:1::3.0 29.6:1::1.3 33.6:1::1.5 11.7:1::2.1 20.1:1::2.4 33.5:1::1.6 13.4:1::3.3 20.4:1::3.4 32.0±2.8 15.8:1::2.2
The linear relationship (~=o.928) between denitrification inhibition and the concentration of ammonium in sediment water at the end of the experiment suggested that dissimilatory nitrate reduction to ammonium increased as the denitrification inhibition increased (Fig. 3), which in tum was related to concentrations of Cd, Cu or Zn (Fig. 3). Further studies are required to identify factors controlling the dissimilatory reduction of nitrate to ammonium and the mineralisation immobilisation ofN in the sediment-water environment using tracers. 3S
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Figure3. Relationship betweendenitrification inhibitionand the concentration of ammoniumin sedimentwater receiving Cd, Cu and Zn.
GENERAL DISCUSSION AND CONCLUSION Even though the results from this laboratory study, in which relatively high levels of Cd, Cu and Zn were used, are difficult to directly extrapolate to field situations, they showed that denitrification inhibition due to Cd, Cu or Zn accumulation is likely to occur in wetland sediments receiving high levels of Cd, Cu or Zn through biosolids leaching and recycled water application. Environmental factors such as the sediment parent material, rainfall, temperature, vegetation and other wetland management practices influence the bioavailability of Cd, Cu or Zn to denitrifying organisms and N20 release in the field. The study showed that
Impact of heavy metals on denitrification
355
addition of Cd, Cu or Zn inhibited denitrification in the sediment-water environment and the inhibition is concentration and time dependent. As the concentration of Cd, Cu or Zn increased, the denitrification inhibition increased but with time, denitrification inhibition decreased for a given metal and concentration. For a given metal concentration the largest inhibition occurred with Cd followed by Zn or Cu. It was also evident from the study that dissimilatory reduction of nitrate to ammonium may also have occurred as an alternative denitrification pathway. The formation of ammonium depended upon the type of the metal and its concentration. As the concentration of metal increased, the denitrification inhibition increased and the ammonium concentration in the sediment water increased. For a given metal concentration the largest ammonium formation was observed with Cd followed by Zn or Cu. The study suggested that the affinity of Cu to sediment was greater than Zn or Cd and resulted in less availability in sediment solution (Ross, 1994). Wetland sediments are complex substrates and completely different levels of the availability occur for different heavy metals, depending upon the ionic size of the metal, sediment solution chemistry and the organic matter and clay content of the sediment. These characteristics determine the effective contact between heavy metals and microorganisms and may result in a relatively high metal load tolerance before toxic effects are produced. However, once the sediment becomes saturated, small changes in metals concentration may be able to cause significant increases in toxicity, as the bioavailable concentration would increase directly with the amount added. This could happen in sewage or runoff treating wetlands which received elevated loads of heavy metals. In addition, the microbial tolerance to toxicants is not unlimited. It is thus necessary to consistently monitor and regulate the loads of heavy metals and other toxicants at a sustainable level to maintain effective N removal in natural or constructed wetland systems receiving recycled or runoff water. REFERENCES APHA. (1992). Standard Methods for Examination of Water and Wastewater. 18th edition. Am. Pub. Health Assoc., Washington. D.C. Bavor, H. J. and Mitchell, D. S. (Eds.) (1994). Wetland systems in water pollution control. Wat. Sci. Tech. 29(4). Bergeron. V, Blais, J. S., Wharf, V. and Marshall, W. D. (1993). Toxicity of tributyltin chloride to anaerobic nitrogen transformations in sediment and porewater. J. Environ. Qual., 22, 528-536. Buresh, R. J. and Patrick, W. H. (1978). Nitrate reduction to ammonium in anaerobic soil. Soil Sci. Soc. Am. J., 42, 913-918. Caskey, W. H and Tiedje, J. M. (1979). Evidence for Clostridia as agents of dissimilatory reduction of nitrate to ammonium in soils. Soil Sci. Soc. Am. J., 43, 931-936. Freeze, R. A. and Cherry, J. A. (1979). Groundwater. Prentice Hall, Englewood Cliffs, N.J. Gilliam, J. W. (1994). Riparian wetlands and water quality. J. Environ. Qual., 23, 896-900. Harris, R. F. (1982). Energetics of nitrogen transformations. In: Nitrogen in Agricultural Soils. F. J. Stevenson (Ed.], pp. 833·890. American Society of Agronomy, Madison, Wisconsin. Heaton. T.H.E. (1986). Isotopic studies of nitrogen pollution in the hydrosphere and atmosphere: A review. Chem. Geol. 59, 87. Howard, S. P., Rowe, D. Rand Tchobanoglous, G. (1985). Advanced wastewater treatment. McGraw-Hili, Inc. Singapore. pp. 294-299. Korom, S. F. (1992). Natural denitrification in the saturated zone: A review. Water Resour. Res., 28, 1657-1668. Magee, P. N. (1982). Nitrogen as a potential health hazard Phil. Trans. R. Soc. London Ser. B., 296, 543-550. McGrath, P. S., Brooks, P. C. and Giller, K. E. (1988). Effects of potentially toxic metals in soils derived from past applications of sewage sludge on nitrogen fixation by Trifolium repens. Soil Bioi. Biochem., 20,415-424. Miller, R. W. and Donahue, R. L. (1990). An introduction to soils and plant growth, Prentice-Hall International. Sixth Edition. pp.537-575. Premi, P. R and Cornfield, A. H. (1969). Effect of addition ofCn, Mn, Zn and Cr compounds on ammonification and nitrification during incubation of soil. Plant Soil.• 31, 345-352. Roser, D. 1. and Baver, H. 1. (1994). SWAMp™. A computerised decision support system for employing wetlands in the biological removal of nutrients and other water pollutants. In: Biological Nutrient Removal 2. Albury, Australia, pp. 227232. Ross, S. M. (1994). Detention, transformation and mobility of toxic metals in soils. In: Toxic metals in soil-plant systems. Ross S. M (Ed.). pp. 63-152. John Wiley & Sons Ltd. Sakadevan, K. and Bavor, H. J. (1998). Phosphate adsorption characteristics of soils, slags and zeolite to be used as substrates in constructed wetland systems. Wat. Res, 32, 393-399. Sumner, M. E. and McLaughlan, M. J. (1996). Adverse impacts of agriculture on soil, water and food quality, In: Contaminant and the soil environment in the Australasia and Pacific Region. R. Naidu, R. S. Kookana, D. P. Oliver, S. Rogers and M. J. Mclaughlin. (Eds). pp. 125-81. Kluwer, London. Veda, K., Kobayashi, M and Takahashi, E (1988). Effect of anionic heavy metals on ammonification and nitrification in soil. Soil Sci. Plant Nut., 34,139-146. Zann, L. P, (1995). Our sea, our future, major findings of the state of the marine environment report for Australia. Department of Environment, Sport and Territories, Canberra.
Wal SCI Tech Vol. 40, No 3, pp. 349-355, I
Pergamon
Printed in Great Bntam. All nghts reset 0273-1223/99 $20.00 + I
PH: S0273-1223(99)00471-0
IMPACT OF REAVY METALS ON DENITRIFICATION IN SURFACE WETLAND SEDIMENTS RECEIVING WASTEWATER K. Sakadevan*, Huang Zheng** and H. J. Bavor* • Water Research Laboratory. University of Western Sydney Hawkesbury, Richmond, NSW-2753, Australia •• Institute ofEnvironmental Medicine, Medical University. Wuhan 430030, P.R. China
ABSTRACT Denitrification in sediment-water systems is a predominant process in the removal of nitrogen from wetlands and sediments receiving recycled water. In this study the impact ofcadmium (Cd), copper (Cu) and zinc (Zn) on denitrification was examined for a wetland sediment receiving recycled water. Results from the study showed that application of 100 mg Cd kg" sediment had no effect (P<0.05) on denitrification (161±2.7 mg N kg" sediment) compared to the control (162±2.4 mg N kg" sediment) which did not receive Cd, Cu or Zn. Addition of )00 mg Cu or Zn kg" sediment significantly increased denitrification (170± 1.8 and 168±2.7 mg NzO-N kg" sediment for Cu and Zn, respectively) compared to the control treatment. Addition of Cd, Cu or Zn at 500 or 1000 mg kg" sediment significantly decreased (P<0.05) total denitrification compared to the control and treatments. which received 100 mg Cd, Cu or Zn kg" sediment. For a given heavy metal concentration the largest denitrification inhibition occurred with Cd (30.9%) followed by Zn (24.9%) and Cu (18.9%) over a period of seven days. The amount of ammonium in the sediment water increased 1D all treatments receiving Cd, Cu or Zn and the concentration increased as the concentration of Cd, Cu or Zn increased in the sediment-water environment. For a given heavy metal concentration, the largest increase in ammonium occurred in treatments receiving Cd (31.1±0.9 mg N kg" sediment) followed by Zn (24.8±0.5 mg N kg" sediment) and Cu (l7.0±0.3 mg N kg" sediment). Denitrification inhibition was linearly related to the concentration of ammonium in sediment water (r=0.928). In general, the study showed that the addition of Cd, Cu or Zn inhibited denitrification and increased the concentration of ammonium in the sediment-water environment. iCI 1999 Published by Elsevier Science Ltd on behalf of the IAWQ. All nghts reserved
KEYWORDS Denitrification; heavy metal; inhibition; recycled water; sediment; run-off water; wetlands. INTRODUCTION Constructed wetland treatments of recycled and runoff waters are currently being practiced in man) countries as an alternative to the direct disposal of recycled or runoff water to coastal and inland waters (Bavor and Mitchell, 1994; Roser and Bavor, 1994; Sakadevan and Bavor, 1998). Recycled water and agricultural and urban community storm water runoff contain high concentrations of nitrogen (N) in the forms of oxidised nitrogen, ammonium and organic N. The excess concentration of N along with P ir recycled and run-off waters may cause eutrophication of surface waters leading to accelerated plant growth, changes in plant species composition and decrease the dissolved oxygen concentration of water (Zann 349
K. SAKADEVAN et al.
350
1995). Nitrate is also one of the most common contaminants in ground water and high concentrations of nitrate have been recognised as a health hazard in drinking water (Freeze and Cherry, 1979; Magee, 1982). When recycled or run-off water is applied to wetland systems, the ammonium and organic N are converted to nitrate through mineralisation and nitrification. Not being strongly adsorbed by sediment, nitrate may move to surface and ground waters. To protect ground and surface waters, the excess nitrate should be minimised via denitrification processes. The nitrification-denitrification process in sediment-water systems is considered to be a predominant process in the removal ofN from sediment-water systems (Korom, 1992). Denitrification in surface sediments receiving recycled water can prevent the movement of nitrate to ground and surface waters (Gilliam, 1994). In this biological process, ammonium is first oxidised to nitrate and then sequentially reduced to molecular N by denitrifying organisms. The process may be inhibited by sustained loads of various potentially toxic compounds including heavy metals, pesticide residues and other organic compounds such as polychlorinated biphenyls in recycled water, which are toxic to denitrifying organisms. An accumulation of heavy metals by long term application of biosolids may decrease the activity of sediment organisms (McGrath et al., 1988). Some effluent and biosolids contain metal concentrations up to 3,400 mg Cd, 10,400 mg Cu and 27,800 mg Zn kg'! dry weight biosolid (Miller and Donahue, 1990), which when applied to sediment may result in accumulation of large amounts of Cd, Cu and Zn in the sediment and may affect denitrification. The objective of the present study is to examine the effect of cadmium (Cd), copper (Cu) and zinc (Zn) and their concentrations on denitrification rates in wetland sediments and its relationship to N removal processes in the treatment of recycled water. MATERIALS AND METHODS Sediment The sediment used in the study was collected from a wetland drain on the University of Western Sydney Hawkesbury Dairy-Farm, situated 60 km northwest of Sydney, NSW, Australia. The farm is currently irrigated with recycled water obtained from the Richmond Sewage Treatment Plant, situated very close to the farm, and the wetland drain receives runoff from the irrigation. Sediment samples (50 cores of the 150 mm surface sediment, within a transect of 50 m x 10m) were collected from one wetland drain of the irrigated farm. All 50 cores of field moist sediment samples were bulked and sieved to less than 2 mm and stored at 4°C. Some chemical characteristics of the sediment are given in Table 1. Table 1. Chemical characteristics ofthe sediment Total C (g C kg,l sediment) Total N (g N kg,l sediment) Total P (mg P kg,lsediment) pH (l :2.5 H20) Extractable nitrate (mg N kg" sediment) Extractable ammonium ( mg kg" sediment) Exchangeable calcium (C mole(+) kg" sediment) Exchangeable magnesium (C mole(+~ kg') sediment) Exchangeable sodium (C mole(+)kg' sediment) Sodium Adsorption Ratio Total Cd (mg Cd kg" sediment) Total Cu (mg Cu kg" sediment) Total Zn (mg Zn kg"\ sediment)
16.6 1.4 603 6.3 4.8 3.1 2.55 1.83 0.91 0.62 1.2 4.8 17.8
Incubation experiment Field moist sediment samples (equivalent to 10 g dry weight) were weighed into 120 ml serum vials and the vials were treated separately with 1 ml of glucose solution (carbon source) containing 1800 g C L'I, 1 ml of heavy metal solution (separate heavy metal solutions for different treatments) and 7 ml deionised water. Including control, a total of 10 treatments were established and three replications were used for each treatment. The individual vials were then closed with rubber suba seal serum caps. A gentle stream of N2
Impact of heavy metals on denitrification
351
(300 ml minute") was directed over the slurry for three minutes by using an inlet and outlet needle, which pierced through the septum. After directing N2 over the slurry, 5 ml of head-space N2 from the vial was replaced with 5 ml of acetylene (C2H2) using a gas-tight syringe. Finally, 1 ml of sodium nitrate containing 1800 mg N L was injected to the vial to give a final concentration of 180 mg N Lo 1 in solution or 180 mg N kg" sediment. The final C concentration in solution also was 180 mg C LO\ or 180mg C kg" sediment. The control treatment was identical to other treatment except it received 1 ml of water instead of heavy metal solution. The vials were gently shaken for one hour and then incubated at room temperature. O
'
,
Sediment, water and gaseous measurement The total C content of the sediment was measured before the experiment using a carbon-nitrogen-sulphur analyser (Fison, NA 1500 NCS). The total N content of the sediment was measured using an Alpkem Flow Solution 111 flow injection analyser (APHA, 1992) after digesting the sediment with a mixture of sulphuric acid, potassium sulphate and selenium (Kjeldahl mixture). The mineral N (nitrate + ammonium) in the sediment was measured using an Alpkem Flow Solution III system as above, after extracting the sediment with 2M KCl (1:10 ratio). After the experiment, the sediment water was filtered and the mineral N (nitrate + ammonium) concentration of the filtrate was measured as above. The sediment denitrification was estimated by the acetylene inhibition technique to yield a potential denitrification rate. At 24-hour intervals, a 0.2 ml headspace sample was removed from each vial using a gas-tight syringe and nitrous oxide (N20) emission from the sediment-water system was analysed by gas chromatography (GC) with a thermal conductivity detector. The denitrification inhibition rate was calculated as follows:
Statistical analysis Data were analysed using the analysis of variance (ANOVA) of SysStat using heavy metal type and concentration as the main effects on denitrification rate and ammonium and nitrate concentrations in sediment water at the end of the experiment. The interaction between three heavy metal concentrations on denitrification and ammonium formation was examined. Differences between the means within each effect were evaluated using a Fisher's protected least-significant-difference test. RESULTS AND DISCUSSION Impacts of Cd. Cu and Zn on denitrification The cumulative amount of N20 released with time for all treatments showed that significant (P
352
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7
8
Time (days) Figure I. Cumulative amount ofN20 oxide produced for sediment receiving (a) cadmium; (b) copper; and (c) zinc at no metal (-8-) and addition of 100 (--A-), 500 ( ) and 1,000 (__) mg metal kg" sediment.
The decrease in N20 released was greater (P
Impact of heavy metals on denitrification
353
than for Cu or Zn for a given concentration. Copper and Zn had similar effects on denitrification and N20 release from sediment-water which received 100 and 500 mg Cu or Zn kg'! sediment, but the decrease in N20 was greater for treatment which received 1000 mg Zn kg'! sediment than treatment that received 1000 mg Cu kg" sediment. Inhibitory effects were observed at 500 and 1000 mg Cd, Cu and Zn kg" sediment amendments (Fig. 2), and indicated that the order of inhibition was Cd>Zn>Cu for all levels of addition.
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'"
20
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10 0 0
2
4 Time (days)
6
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Figure 2. Denitrification inhibition in sediment receiving 500 ( ) and 1000 (~) mg Cd kg" sediment, 500 (~) and 1000 (-e-) mg Cu kg" sediment and 500 (-9-) and 1000 (-+-) mg Zn kg' sediment.
The denitrification inhibition decreased during the incubation period (Fig. 2). This decrease may be due to: (1) among the wide range of denitrifying micro-organisms, at least a portion of the original population was able to overcome the inhibitory effects and developed tolerance to Cd, Cu or Zn. Similar phenomenon was observed that adaptation to heavy metals and selection of the resistant strains could occur rapidly in the sediments among a variety of nitrifying micro-organisms (Rother et al., 1982) and denitrifying population was also capable of recovering from the inhibitory effects ofthe toxicants (Bergeron et al., 1993); and/or (2) the amount of the bioavailable Cd, Cu or Zn decreased in the sediment-water due to adsorption, chelation and precipitation processes in the sediment (Ross, 1994). Ammonium and nitrate in sediment solution The concentration of ammonium in sediment solution in all treatments including the control was increased at the end of the experiment (Table 2), while nitrate concentrations decreased in all treatments (including control), even though about 180 mg nitrate N kg" sediment was added to the sediment at the beginning of the experiment (Table 2). The increases in ammonium concentrations were significantly greater in treatments which received 500 or 1000 mg Cd, Cu or Zn kg'! sediment compared to the control (P<0.05) and treatments which received 100 mg kg'! sediment of Cd, Cu or Zn. There was no significant difference in the amount of ammonium in sediment solution between control and treatments amended with 100 mg Cd, Cu or Zn kg" sediment (Table 2). The largest ammonium accumulation was observed in treatments which received 1000 mg Cd, Cu or Zn kg'! sediment. For a given concentration, the largest ammonium accumulation was observed for Cd followed by Zn and Cu (Table 2). With the nitrate reduction to N 2 0 and N 2 by denitrification processes, dissimilatory reduction of nitrate to ammonium N could have occurred in the same anoxic and carbon rich sediment-water environment, mediated by spore forming genera of Clostridium and Bacillus (Buresh and Patrick, 1978; Caskey and Tiedje, 1979). Both denitrification and the dissimilatory nitrate reduction are processes in which they make energy available to the cell for growth and maintenance (Harris, 1982). The denitrification inhibition by Cd, Cu or Zn may have resulted in some nitrate diversion into the pathway of reduction to ammonium rather tpan N20. This would result in more ammonium accumulation in sediment systems containing 500 or 1000 mg Cd, Cu or Zn kg'! sediment than control and treatments which received 100 mg Cd, Cu or Zn kg" sediment (Table 2). It was observed in the study that all added nitrate had not been denitrified (lower N 20 measured than nitrate added to the sediment-water environment) in sediment-water system of all treatments. Mass balance calculations of nitrate and
K. SAKADEVAN et al.
354
ammonium showed that the mineral N (sum of nitrate and ammonium) concentrations in sediment solutions at the end of the experiment had decreased for all treatments, reflecting a net loss of ammonium and/or nitrate from sediment solution (Table 2). This "loss" from sediment solution is hypothesised to have occurred through ammonia volatilisation or conversion to organic N by microbial biomass uptake. Table 2. Concentration of ammonium and nitrate in sediment water at the end of the experiment for all treatments Net mineral N lost Concentration Ammonium Nitrate Metal from sediment solution mg N kg" sediment
Cu
Zn
Control
7.8:1::0.85 17.1:1::2.43 31.1:1::0.72 6.1:1::0.70 11.5:1::1.08 17.0±0.32 5.8:1::0.90 15.1:1::1.42 24.8:1::0.52 9.6:1::0.82
100 500 1000 100 500 1000 100 500 1000
Cd
No added metal
1.1:1::0.15 1.5:1::0.36 1.5:1::0.06 0.9±0.4 1.2:1::0.06 0.8:1::0.06 1.0±.11 0.9±0.06 1.0±0.21 0.8:1::0.06
17.8:1::3.0 29.6:1::1.3 33.6:1::1.5 11.7:1::2.1 20.1:1::2.4 33.5:1::1.6 13.4:1::3.3 20.4:1::3.4 32.0±2.8 15.8:1::2.2
The linear relationship (~=o.928) between denitrification inhibition and the concentration of ammonium in sediment water at the end of the experiment suggested that dissimilatory nitrate reduction to ammonium increased as the denitrification inhibition increased (Fig. 3), which in tum was related to concentrations of Cd, Cu or Zn (Fig. 3). Further studies are required to identify factors controlling the dissimilatory reduction of nitrate to ammonium and the mineralisation immobilisation ofN in the sediment-water environment using tracers. 3S
30
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•
2
r - O.928
S
O+---f--+--lf---+---I---+---f
o
S
10
IS 20 e;. Inhibition
2S
30
3S
Figure3. Relationship betweendenitrification inhibitionand the concentration of ammoniumin sedimentwater receiving Cd, Cu and Zn.
GENERAL DISCUSSION AND CONCLUSION Even though the results from this laboratory study, in which relatively high levels of Cd, Cu and Zn were used, are difficult to directly extrapolate to field situations, they showed that denitrification inhibition due to Cd, Cu or Zn accumulation is likely to occur in wetland sediments receiving high levels of Cd, Cu or Zn through biosolids leaching and recycled water application. Environmental factors such as the sediment parent material, rainfall, temperature, vegetation and other wetland management practices influence the bioavailability of Cd, Cu or Zn to denitrifying organisms and N20 release in the field. The study showed that
Impact of heavy metals on denitrification
355
addition of Cd, Cu or Zn inhibited denitrification in the sediment-water environment and the inhibition is concentration and time dependent. As the concentration of Cd, Cu or Zn increased, the denitrification inhibition increased but with time, denitrification inhibition decreased for a given metal and concentration. For a given metal concentration the largest inhibition occurred with Cd followed by Zn or Cu. It was also evident from the study that dissimilatory reduction of nitrate to ammonium may also have occurred as an alternative denitrification pathway. The formation of ammonium depended upon the type of the metal and its concentration. As the concentration of metal increased, the denitrification inhibition increased and the ammonium concentration in the sediment water increased. For a given metal concentration the largest ammonium formation was observed with Cd followed by Zn or Cu. The study suggested that the affinity of Cu to sediment was greater than Zn or Cd and resulted in less availability in sediment solution (Ross, 1994). Wetland sediments are complex substrates and completely different levels of the availability occur for different heavy metals, depending upon the ionic size of the metal, sediment solution chemistry and the organic matter and clay content of the sediment. These characteristics determine the effective contact between heavy metals and microorganisms and may result in a relatively high metal load tolerance before toxic effects are produced. However, once the sediment becomes saturated, small changes in metals concentration may be able to cause significant increases in toxicity, as the bioavailable concentration would increase directly with the amount added. This could happen in sewage or runoff treating wetlands which received elevated loads of heavy metals. In addition, the microbial tolerance to toxicants is not unlimited. It is thus necessary to consistently monitor and regulate the loads of heavy metals and other toxicants at a sustainable level to maintain effective N removal in natural or constructed wetland systems receiving recycled or runoff water. REFERENCES APHA. (1992). Standard Methods for Examination of Water and Wastewater. 18th edition. Am. Pub. Health Assoc., Washington. D.C. Bavor, H. J. and Mitchell, D. S. (Eds.) (1994). Wetland systems in water pollution control. Wat. Sci. Tech. 29(4). Bergeron. V, Blais, J. S., Wharf, V. and Marshall, W. D. (1993). Toxicity of tributyltin chloride to anaerobic nitrogen transformations in sediment and porewater. J. Environ. Qual., 22, 528-536. Buresh, R. J. and Patrick, W. H. (1978). Nitrate reduction to ammonium in anaerobic soil. Soil Sci. Soc. Am. J., 42, 913-918. Caskey, W. H and Tiedje, J. M. (1979). Evidence for Clostridia as agents of dissimilatory reduction of nitrate to ammonium in soils. Soil Sci. Soc. Am. J., 43, 931-936. Freeze, R. A. and Cherry, J. A. (1979). Groundwater. Prentice Hall, Englewood Cliffs, N.J. Gilliam, J. W. (1994). Riparian wetlands and water quality. J. Environ. Qual., 23, 896-900. Harris, R. F. (1982). Energetics of nitrogen transformations. In: Nitrogen in Agricultural Soils. F. J. Stevenson (Ed.], pp. 833·890. American Society of Agronomy, Madison, Wisconsin. Heaton. T.H.E. (1986). Isotopic studies of nitrogen pollution in the hydrosphere and atmosphere: A review. Chem. Geol. 59, 87. Howard, S. P., Rowe, D. Rand Tchobanoglous, G. (1985). Advanced wastewater treatment. McGraw-Hili, Inc. Singapore. pp. 294-299. Korom, S. F. (1992). Natural denitrification in the saturated zone: A review. Water Resour. Res., 28, 1657-1668. Magee, P. N. (1982). Nitrogen as a potential health hazard Phil. Trans. R. Soc. London Ser. B., 296, 543-550. McGrath, P. S., Brooks, P. C. and Giller, K. E. (1988). Effects of potentially toxic metals in soils derived from past applications of sewage sludge on nitrogen fixation by Trifolium repens. Soil Bioi. Biochem., 20,415-424. Miller, R. W. and Donahue, R. L. (1990). An introduction to soils and plant growth, Prentice-Hall International. Sixth Edition. pp.537-575. Premi, P. R and Cornfield, A. H. (1969). Effect of addition ofCn, Mn, Zn and Cr compounds on ammonification and nitrification during incubation of soil. Plant Soil.• 31, 345-352. Roser, D. 1. and Baver, H. 1. (1994). SWAMp™. A computerised decision support system for employing wetlands in the biological removal of nutrients and other water pollutants. In: Biological Nutrient Removal 2. Albury, Australia, pp. 227232. Ross, S. M. (1994). Detention, transformation and mobility of toxic metals in soils. In: Toxic metals in soil-plant systems. Ross S. M (Ed.). pp. 63-152. John Wiley & Sons Ltd. Sakadevan, K. and Bavor, H. J. (1998). Phosphate adsorption characteristics of soils, slags and zeolite to be used as substrates in constructed wetland systems. Wat. Res, 32, 393-399. Sumner, M. E. and McLaughlan, M. J. (1996). Adverse impacts of agriculture on soil, water and food quality, In: Contaminant and the soil environment in the Australasia and Pacific Region. R. Naidu, R. S. Kookana, D. P. Oliver, S. Rogers and M. J. Mclaughlin. (Eds). pp. 125-81. Kluwer, London. Veda, K., Kobayashi, M and Takahashi, E (1988). Effect of anionic heavy metals on ammonification and nitrification in soil. Soil Sci. Plant Nut., 34,139-146. Zann, L. P, (1995). Our sea, our future, major findings of the state of the marine environment report for Australia. Department of Environment, Sport and Territories, Canberra.