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Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate
STOTEN-20975; No of Pages 8 Science of the Total Environment xxx (2016) xxx–xxx
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Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate Xiaoning Liu a,b, Yi Tao a,b,⁎, Kuiyu Zhou a, Qiqi Zhang a, Guangyao Chen a, Xihui Zhang a,b a b
Key Laboratory of Microorganism Application and Risk Control of Shenzhen, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China Tsinghua-Kangda Research Institute of Environmental Nano-Engineering & Technology, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China
H I G H L I G H T S
G R A P H I C A L
A B S T R A C T
• AVS in the sediments decreased significantly with the addition of calcium nitrate. • Calcium nitrate contributed to the reduction of phosphate in the water column. • Water quality had effect on the interstitial nitrate and sulfate in the sediment.
a r t i c l e
i n f o
Article history: Received 18 July 2016 Received in revised form 17 September 2016 Accepted 17 September 2016 Available online xxxx Editor: Jay Gan Keywords: Calcium nitrate Sediment Water quality Acid volatile sulphide Phosphorus Total organic carbon
a b s t r a c t In situ sediment remediation technique is commonly used to control the release of pollutants from sediment. Addition of calcium nitrate to sediment has been applied to control the release of phosphorus from sediments. In this study, laboratory experiments were conducted to investigate the effect of water quality improvement on the remediation of river sediment with the addition of calcium nitrate. The results demonstrated that the redox-potential of sediments increased from −282 mV to −130 mV after 28 days of calcium nitrate treatment. The acid volatile sulphide in the sediments significantly decreased (by 54.9% to 57.1%), whereas the total organic carbon decreased by 9.7% to 10.2%. However, the difference between these and water quality improvement was not significant. Due to the addition of calcium nitrate, low phosphate concentration in the water column and interstitial phosphate in the sediment were observed, indicating that the calcium nitrate was beneficial to controlling the release of phosphorus from river sediment. The decrease in phosphorus release could be attributed to the fixation of iron–phosphorus and calcium–phosphorus due to the addition of calcium nitrate. The addition of calcium nitrate to sediment caused the oxidation of sulphide to sulphate, hence resulting in high nitrate and sulphate concentrations in the water column, and high interstitial nitrate and sulphate concentrations in the sediment. The results also showed that only the water quality improvement had a significant effect on the interstitial nitrate and sulphate concentrations in the sediment. © 2016 Elsevier B.V. All rights reserved.
1. Introduction ⁎ Corresponding author at: Graduate School at Shenzhen, Tsinghua University, China. E-mail address: [email protected] (Y. Tao).
Rivers have been the receivers of domestic wastewater effluent, rainwater runoff, agricultural runoff and industrial wastewater for
http://dx.doi.org/10.1016/j.scitotenv.2016.09.149 0048-9697/© 2016 Elsevier B.V. All rights reserved.
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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years, and this has seriously damaged the river water quality (Scholes et al., 2008). Due to rapid increases in economy and population in China, the quantity of pollutants discharged into the rivers has also increased. The continuing development of pollution in water bodies causes them to become turbid and odourous after their carrying capacity of these pollutants has been exceeded. Water pollution mostly occurs in water bodies, which either have a low flow rate or are static. Therefore, water pollution of urban rivers is becoming one of the greatest threats to the aquatic ecosystem in China. Large amounts of organic pollutants from wastewater are accumulated in river sediment, while excessive degradation of the organic matter consumes much dissolved oxygen (DO) resulting in hypoxia of the sediment. Low DO causes a low redox potential [oxidation reduction potential (ORP)], which leads to microbial reduction of sulphate (SO2− 4 ) to hydrogen sulphide (H2S) in the sediment (Berner, 1970). A decrease in the ORP of sediment can also cause the reduction of Fe and Mn, which results in the release of phosphorus (P) adsorbed on Fe and Mn oxide–hydroxide particles (Caraco et al., 1989; Schauser et al., 2006). Subsequently, Fe(II) in the water column and sediment reacts with H2S to form iron sulphide (FeS) because of its low-solubility product and gets adsorbed on the organic particles, causing the water and sediment layer to turn black-coloured. Furthermore, toxic substances, such as ammonia, H2S and other odourous gaseous sulphur compounds, are produced as a result of anaerobic respiration, which leads to odourous phenomenon along the rivers. Water quality can be improved with the mitigation of external pollution sources and advanced treatment of wastewater. However, sediments can continuously release pollutants to the water column, thus delaying the recovery of eutrophic environments (Søndergaard et al., 2007). Previous research indicated that eutrophic ecosystems could be recovered using in situ technologies, such as addition of nitrate, Al, Fe and Ca to the sediment (Ripl, 1976; Foy, 1986; Cooke et al., 1993; Kleeberg et al., 2000; Xu et al., 2008; Yin and Kong, 2015). The addition of these compounds to sediments can improve the oxidation conditions and accelerate the denitrification activity (Shimizu and Nakano, 2009). Ripl (1976) found that the biochemical oxidation could be achieved for remediation purposes by injecting calcium nitrate into the sediment to control P release from the sediment, which would limit the growth of algae in the water body. Apart from controlling the P release from sediment, nitrate is also used to abate the odours caused by sulphate-reducing bacteria in wastewater (Jenneman et al., 1986; Jefferson et al., 2002; Jiang et al., 2009). When nitrate is metabolised, a portion of the organic matter is used to to S2− (Jenneman et al., 1986). The acid volatile sulphide reduce SO2− 4 (AVS) concept is defined as the sedimentary S, which is extracted by 1 mol·L−1 HCl solution, including porewater sulphides and metastable FeS (Berner, 1964). The AVS content in the sediment is strongly correlated with the colour intensity of the sediment, which could be used to assess the pollution of organically enriched sediment (Wilson and Vopel, 2012). Sulphur-containing odourous compounds also originate from AVS in the sediment, and to some extent, AVS can be considered as the source of sulphur-containing odour and potential production of H2S. Thus, reducing AVS content in the sediment and controlling the release of sulphur-related odour to the water column are meaningful and urgent research tasks. P loading is considered to have a key role in the formation of algal blooms and eutrophication (Yin et al., 2013). Since P release from anaerobic sediments usually makes up the major fraction of total P (TP) load in lakes or rivers, therefore the reduction of an external P load by controlling the influent water quality may be insufficient (Cooke et al., 1993). Consequently, a remediation planning to control the P sources should be a priority. Many amendment technologies, such as modified zeolite (Yang et al., 2014), magnetic microparticles (Funes et al., 2016) and thermally modified calcium-rich attapulgite (Yin et al., 2013; Yin and Kong, 2015) have been processed to immobilize both the soluble reactive P in porewater and mobile P in sediment.
Upgradation of wastewater treatment plants and the mitigation of external pollution sources have been considered effective in improving the lake and river water quality (Jeppesen et al., 2007; Fulton et al., 2015). However, both the improvement and remediation of polluted sediment are far slower than that of the water column. Meanwhile, sediment would continue to be an important internal source of nutrients to the water column for a long time, which affects the water quality of overlying water due to a larger nutrient concentration in the sediment porewater than in the water column (Liu et al., 2014). In the present study, the effect of water quality improvement on sediment remediation has been investigated, which is related to the variation in AVS due to the addition of calcium nitrate. 2. Material and methods 2.1. Water and sediment sampling The sediment and water samples used in this study were obtained from the Shajing River (22°46′3″, 113°49′52″) in Shenzhen City, Guangdong, China. This river is a tributary of the Maozhou River, which flows into the Pearl River Estuary. In the present study, the water and sediment samples were obtained from upstream, midstream and downstream of the Shajing River, and were mixed to present the characteristics of water and sediment. The river flows slowly at the depth of 1.5–2.0 m, and receives pollutants of rainfall runoff and occasional industrial and urban wastewaters from the riverside. The water samples were obtained from the middle zone of the river using a glass water sampler, while the sediment samples were collected from the top 20 cm layer of three sites using a piston column sediment sampler (purchased from Nanjing Institute of Geography and Limnology, Chinese Academy of Sciences, China). The samples were homogenized and sieved to remove detritus, plant residues and plastic products. Then, the collected samples were kept cool in a box containing ice cake. These were transferred to the laboratory and stored at a temperature of approximately 4 °C until further use. The water quality and sediment's characteristics of the sampling station are summarized in Tables 1 and 2. 2.2. Experimental design The experiments were conducted using 1 L wide-mouth jars having screw top lids. Before the experiments, the jars were placed in a sterilizing pot at 121 °C for 30 min. The sediment (250 g) and water column (800 mL) were placed in the jars, and subsequent treatment was performed in triplicate. The heights of sediment and overlying water were 5 and 10 cm, respectively. Four treatments were conducted. The first treatment did not receive calcium nitrate (control treatment). The sediments in the other three treatments were given the same dosage of calcium nitrate at three different water column compositions as shown in Table 3. The water column was replaced with pure water, which represented the treated water. Based upon the study of Chen et al. (2013), the application dosage was 1 g of calcium nitrate per 1 kg of wet sediment. Nitrogen gas was injected into the jars when the calcium nitrate was mixed with the sediment to avoid sediment oxidation. The jars were kept closed during the experiments, which were performed at 25 °C in an incubator. The water samples were withdrawn on 0th, 1st, 3rd, 6th, 10th, 15th, 21st and 28th day of the experiment, while those of sediment were withdrawn on 0th and 28th day of the Table 1 Characteristics of the Shajing River water investigated in this study. Unit
pH
DO mg·L
Mean Std
7.26 0.05
0.26 0.04
Eh −1
NH+ 4 -N
mV
mg·L
−68.6 12.2
14.6 1.2
−1
NO− 3 -N mg·L 0.10 0.02
−1
TN mg·L 18.2 0.86
−1
PO3− 4 -P
COD
mg·L−1
mg·L−1
6.2 1.2
84.5 15.0
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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Table 2 Characteristics of the sediments obtained from the Shajing River. Sediment
pH
Unit Mean Std
7.25 0.12
Eh
Water content
Total organic carbon
Acid volatile sulphide
mV
%
mg·g−1
mg S·kg−1
−282 35
78.2 3.2
90.4 3.4
5136 102
experiment. The measured values in the sediment before the experiment (0th day) are defined as “start” values. Nitrogen gas was injected into the flask to drive out oxygen prior to shaking. The diluted slurry was centrifuged (7000 rpm·min− 1), while the supernatant was − filtrated through 0.45 μm Millipore filters for measuring NH+ 4 -N, NO3 2− -P and SO . N, PO3− 4 4 2.3. Measurement method and statistical analysis The water samples were filtered through 0.45 μm Millipore filters before chemical analysis to determine the dissolved nutrients. The pH and DO concentrations of the water column were measured using a Hach HQ40d Dual-input Multi-parameter Meter. According to the standard methods described by the American Public Health Association − 3− 2− (1998), the NH+ 4 -N, NO3 -N, total nitrogen (TN), PO4 -P and SO4 were analyzed by colorimetric analysis by using a spectrophotometer. The pH and ORP of the sediment were measured with a Hach HQ40d Dual-input Multi-parameter Meter at a depth of 3 cm from the surface layer on 0th and 28th day of the experiment. The dry weight was measured by drying the sediment at 105 °C for 24 h, while the water content was calculated based upon the weight loss. The interstitial ions in the sediments were extracted by diluting the sediment sample in 1:4 ratios using pure water. The diluted sample was stirred for 4 h at 200 rpm·min−1. Nitrogen gas was injected into a closed conical flask to maintain the hypoxia condition before shaking it in the shaker. The AVS in the sediment was analyzed using the method described by Hsieh and Shieh (1997). The TOC in the sediment was measured using potassium dichromate titration method (Schumacher, 2002). The fractional composition of TP (different P binding forms) in the sediment was examined using the fractionation scheme of Psenner et al. (1984). This sequential extraction method divides the TP of sediment into pools as shown in Table 4. The amount of P in each fraction is determined as TP described above. The difference between TP and the sum of the extracted pools is the residual P (Pres), which consists mainly of organic bound P. The average values and standard deviations of three replicate samples were calculated. Statistical analysis (two-way ANOVA) was performed to evaluate the significant differences among the means in the sediment by using SPSS Statistics 19.0. 3. Results 3.1. Parameters of the water column The pH value of the water column for all samples exhibited a transient increasing trend before it began to decline consistently (Fig. 1a). The pH values of the treatments with added calcium nitrate were slightly higher than that in the control on the 28th day. The mean pH value in Table 3 The experimental design and water column composition for the four treatments. Treatments
Experimental design
Water column composition
Control Type 1 Type 2
Sediment Sediment and Ca(NO3)2 Sediment and Ca(NO3)2
Type 3
Sediment and Ca(NO3)2
800 mL of water samples 800 mL of water samples 720 mL of water samples and 80 mL of pure water 600 mL of water samples and 200 mL of pure water
the control was 6.82, while those in types 1, 2 and 3 experiments were 7.08, 7.08 and 7.06, respectively. Neither the addition of calcium nitrate to sediment nor water quality improvement had any obvious effect on the pH of the water column. The DO values of the water column in the control showed no significant variation during the entire experiment. However, those of the treatments exhibited a sharply increasing trend during the first several days. The DO values in types 1 and 2 experiments reached the highest values (2.03 and 2.64 mg·L− 1, respectively) on the 3rd day, while type 3 treatment (2.15 mg·L−1) reached the highest value on the 6th day. These values decreased to a lower value on the 10th day, ending up even lower than 0.5 mg·L−1 at the end of the experiment (Fig. 1b). The addition of calcium nitrate to sediment affected only the DO value to some extent during the first several days, while the water quality improvement had no obvious effect on the DO value. The NO− 3 -N concentration in the control showed an increasing trend, whereas that of the calcium nitrate treatments first increased sharply for a short time period (1 day), decreased for 10 days and then increased again with time (Fig. 2a). The concentrations increased to 8.8, 10.4 and 11.9 mg·L−1 on the 28th day in types 1, 2 and 3 experiments, respectively. Overall, the addition of calcium nitrate increased the NO− 3 -N concentration in the water column, and the NO− 3 -N concentration in type 2 experiment exhibited the highest value (11.9 mg·L− 1). The NH+ 4 -N concentration in the water column increased sharply during the initial stage of the experiment, reaching a maximum of nearly 40 mg·L−1 on the 10th day. Then, the concentration began to decrease consistently (Fig. 2b). The NH+ 4 -N concentrations in the calcium nitrate treatments were slightly higher than that in the control on the 28th day. The TN concentration exhibited a similar trend. It reached the maximum value of 44.0, 47.6, 46.1 and 46.1 mg·L−1 on the 10th day in control, types 1, 2 and 3 experiments. Afterwards, the concentration began to decrease consistently (Fig. 2c). The TN concentration in the water column in the control was slightly lower than those in the treatments. However, all of these reached the maximum values on the 10th day. The PO34 −-P concentration in the water column in the control decreased during the initial period (1st day) and then increased consistently with time (Fig. 2d), reaching a maximum value of 2.4 mg·L−1 on the 10th day. The concentration subsequently exhibited a slowly declining trend and decreased to 2.0 mg·L−1 on the 28th day. On the other hand, the PO3− 4 -P concentrations exhibited a decreasing trend in the calcium nitrate treatments, and the values on the 28th day were found to be 1.2, 0.7 and 0.7 mg·L−1 in types 1, 2 and 3 experiments, respectively. The SO24 − concentration in the water column decreased slowly from 98.5 mg·L−1 to 79.3 mg·L−1 in the control, whereas it exhibited an increasing trend in the calcium nitrate treatments (Fig. 2e), reaching 138.1, 150.8 and 119.2 mg·L−1 on the 28th day in types 1, 2 and 3 experiments, respectively.
Table 4 Sequential extraction methods of different P binding forms in the sediment. Mark
Extraction/Time
Expected P binding forms
NH4Cl-P NaOH-P BD-P H2SO4-P Pres
1.0 M NH4Cl/2 h 0.1 M NaOH/16 h 0.3 M Na2S2O4/NaHCO3/1 h 0.25 M H2SO4/2 h
Pore water P, Loosely adsorbed to surfaces Inorganic P bound to Al Redox-sensitive P Ca-binding P Organic and other refractory P
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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X. Liu et al. / Science of the Total Environment xxx (2016) xxx–xxx 7.6
3.5
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control type 1 type 2 type 3
7.4
(b)
control type 1 type 2 type 3
3.0 2.5
pH
DO (mg·L -1)
7.2
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6.6 0
5
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0
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Time (d)
Fig. 1. Variation in (a): pH and (b): DO of the water column (n = 3).
3.2. Interstitial ions in the sediment The variations in the interstitial ions in the sediment are presented in Fig. 3. The interstitial NO− 3 -N concentration in the sediment in the control increased from 6.1 mg·kg−1 to 7.1 mg·kg− 1. The interstitial NO− 3 -N concentrations increased significantly to values of 22.9, 18.1 and 19.0 mg·kg−1 in types 1, 2 and 3 experiments, respectively (Fig. 3a). The interstitial NO− 3 -N concentrations in the sediment in types 1, 2 and 3 experiments were significantly different from that in the control. Type 1 experiment had the greatest effect on the interstitial NO− 3 N concentration in the sediment, which was also significantly different from those in types 2 and 3 experiments. The interstitial NH+ 4 -N concentration in the sediment of the control decreased from 646.7 mg·kg−1 to 472.7 mg·kg−1, whereas the corresponding values in types 1, 2 and 3 experiments on the 28th day were found to be 406.7, 385.3 and 393.0 mg·kg−1, respectively. The interstitial NH+ 4 -N concentration in the sediment in the control was higher than that in types 1, 2 and 3 experiments (Fig. 3b), and was found to be significantly different from those in types 2 and 3 experiments. The interstitial PO3− 4 -P concentration in the sediment of the control decreased from 2.44 mg·kg− 1 to 2.00 mg·kg−1, whereas the corresponding values in types 1, 2 and 3 experiments were found to be 0.36, 0.27 and 0.41 mg·kg−1, respectively. The interstitial PO3− 4 -P concentration in the sediment in types 1, 2 and 3 experiments were found to be significantly different from that in the control. The intersticoncentration in the sediment of the control increased from tial SO2− 4 787 mg·kg−1 to 813 mg·kg−1, whereas the corresponding values for types 1, 2 and 3 experiments were found to be 1012, 1205 and 1201 concentrations in the sedmg·kg−1, respectively. The interstitial SO2− 4 iments in types 2 and 3 experiments were significantly different from that of the control; however the value was not significantly different from that in type 1 experiment.
3.3. Variations in AVS and TOC in the sediment The AVS content in the control was slightly lower than that in the “start” (before the experiment). Compared with that in the control, the AVS content in the sediments of treatments decreased significantly with the addition of calcium nitrate. The corresponding values in types 1, 2 and 3 experiments decreased from 5136 mg S·kg−1 to 2205, 2313 and 2223 mg S·kg−1, respectively (Fig. 4a). However, no significant difference was observed among the values of three treatments with different water qualities. The variation in TOC of the sediments is presented in Fig. 4b. No obvious difference was observed in the TOC content of the sediment in the control (the value decreased from 90.4 mg·kg− 1 to 88.1 mg·kg− 1). However, the TOC contents significantly decreased
from 90.4 mg·kg−1 to 81.2, 81.6 and 81.2 mg·kg−1 in types 1, 2 and 3 experiments, respectively. Similarly, water quality had no significant effect on the TOC content of the sediment. 3.4. P fraction in the sediment The distributions of different P binding forms after 28 days of the addition of calcium nitrate are presented in Fig. 5. The fraction of P binding forms changed, however water quality had no significant effect on them. No obvious difference was observed in the NH4Cl-P in the sediment of the control. On the other hand, NH4Cl-P decreased from 62.95 mg·kg−1 to 5.32, 5.55 and 5.19 mg·kg−1 for the types 1, 2 and 3 experiments, respectively. In the original sediment, P was mostly NaOH-P and H2SO4-P, while BD-P and NH4Cl-P accounted for only a small fraction. After adding calcium nitrate, the amount of NH4Cl-P in the sediment decreased, whereas the amounts of NaOH-P, H2SO4-P and BD-P increased. Furthermore, Pres exhibited a slightly decreasing trend in all treatments. 4. Discussion 4.1. Effect of water quality Water quality of the river had a certain effect on the interstitial NO− 3 N concentration in the sediment and on the NO− 3 -N concentration in the water column. Unlike in type 1 experiment, higher NO− 3 -N concentration of water column was found in types 2 and 3 experiments, thus indicating their improved water quality compared with that in type 1 experiment. However, water quality had insignificant effect on the in+ terstitial NH+ 4 -N concentration in the sediment and on the NH4 -N and TN concentrations in the water column. Furthermore, water quality had no significant effect on the AVS and TOC in the sediment (Fig. 4). The PO3− 4 -P concentrations in types 2 and 3 experiments on the 28th day were lower than that in type 1 experiment (Fig. 2d). Nevertheless, water quality had no significant effect on the variation of P fraction in the sediment. Similarly, with the improvement of water quality, a lower interstitial PO3− 4 -P in the sediment in types 2 and 3 experiments were observed than that in type 1 experiment, though there was no significant difference between the values. Thus, the variation in water quality had no obvious effect on both the interstitial PO3− 4 -P in the sediment and PO3− 4 -P concentration of the water column. Sediment nutrient concentrations are generally weakly correlated with water quality of the lake (Trolle et al., 2009). This means that the water quality has no obvious effect on the variation in sediment pollutants. However, improvement of water quality theoretically increases the diffusive flux at the sediment–water interface, thus leading to a higher release rate of pollutants from the sediment. The results of variation in several major pollutants of water column did now show such
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
X. Liu et al. / Science of the Total Environment xxx (2016) xxx–xxx 50
20
(a)
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control type 1 type 2 type 3
16
(b)
control type 1 type 2 type 3
40
NH4+-N (mg·L -1)
14
NO3--N (mg·L -1)
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(d)
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PO43--P (mg·L -1)
TN (mg·L -1)
40
30
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2.0
1.5
1.0
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0.5
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10
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Time (d) 180
0
5
10
15
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Time (d)
(e)
150
SO42- (mg·L -1)
120
90
60
control type 1 type 2 type3
30
0 0
5
10
15
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25
30
Time (d) + 3− 2− Fig. 2. Variation in (a): NO− in the water column (n = 3). 3 -N (b): NH4 -N (c): TN (d): PO4 -P and (e): SO4
trend (Fig. 2). Although the addition of calcium nitrate had an obvious effect on the variation in ions in the water column, water quality improvement had no obvious influence on the remediation of river sediment due to the addition of calcium nitrate, particularly on the variation in the AVS content in the sediment. Given the short incubation time (28 days) in this study, a longer time is required to assess the function of water quality improvement in remediating the river sediment. 4.2. Effect of calcium nitrate 4.2.1. Sulphur The AVS content in the sediment significantly decreased by 54.9% to 57.1%, while the TOC in the sediment decreased by 9.7% to 10.2% (Fig. 4). Janke et al. (2011) also found that the AVS content in the sediment decreased by 99% after 135 days. Yamada et al. (2012) found that the AVS content in the sediment, which was treated with calcium nitrate,
decreased by 99%. However, the AVS content in the sediment treated with calcium nitrate, was reduced by only 54.9% to 57.1% (Fig. 4), and exhibited a lower reduction rate than that indicated in the literature (Janke et al., 2011; Yamada et al., 2012). Firstly, the reason might be the different in dosages of calcium nitrate as the denitrification activity is proportional to the nitrogen concentration (Wang et al., 2007). Secondly, the loss of nitrate caused by the high water solubility of calcium nitrate in the sediment water could not be directly measured. The nitrate loss may be estimated by both the increase in NO− 3 -N concentration in the water column and interstitial NO− 3 -N in the sediment. Lastly, the current study used a shorter incubation time (28 days) than that of the Yamada et al. (2012), who used a 145 days incubation time. The AVS content in the sediment is related to seasonal and depth variations of TOC, sediment redox potential (Eh) and sulphate concentration in the sediment cores and porewater (Yin and Fan, 2011). The
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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X. Liu et al. / Science of the Total Environment xxx (2016) xxx–xxx 30
800
(a)
25
a 600
b
20
b
b
NH4+-N (mg·kg -1)
NO3--N (mg·kg -1)
(b)
a
bc c
c
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15
10
c
200
c 5
0
0
start
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control
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start
control
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(d)
(c)
4.0
a
1500
a
ab
a
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a
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SO42- (mg·kg -1)
PO43--P (mg·kg -1)
type 1
b
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2.5 2.0 1.5
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b
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start
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start
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type 1
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− 3− 2− Fig. 3. Comparison of interstitial (a): NH+ in the sediment on the 28th day after various treatments (n = 3). 4 -N (b): NO3 -N (c): PO4 -P and (d): SO4
reduction in AVS content in the sediment was mainly attributed to two reasons. Firstly, the levels of sulphate-reducing bacteria decreased during the prolonged exposure to an oxidizing environment and high N2O concentration (Mahmood et al., 2007). Due to these, the microbial reduction of SO24 − to H2S was inhibited. The ORP of the sediment in types 1, 2 and 3 experiments in this study increased from − 282 mV to −130, −127.4 and −128.2 mV, respectively, thus resulting in the inhibition of sulphate-reducing bacteria. The decrease in the ORP of the sediment was caused by the build-up of produced N2O and NO, while the reduction in the amount of organic matter and the production of N2 during denitrification resulted in an oxidizing environment (Jenneman et al., 1986). Secondly, nitrate acted as an electron acceptor, whereas the sulphide in the sediment was oxidized to SO2− 4 in the presence of nitrate (Ripl, 1976; Jenneman et al., 1986). Jiang et al. (2009) assumed that sulphide was oxidized in two steps under anoxic conditions
6000
after the addition of nitrate. The first step was the oxidation of sulphide to S, while the second step was further oxidation of elemental sulphur to SO2− 4 . The two steps occurred sequentially with the second step occurring primarily after sulphide depletion, whereas the first step was considerably faster than the second step. This inference could be supported by both the increase in SO2− 4 concentration in the water column in the sediment (Fig. 3d). Janke et al. (Fig. 2e) and the interstitial SO2− 4 (2011) also stated that the oxidation of sulphide would increase both concentration and the availability of Fe and Mn, which were the SO2− 4 previously complexed as FeS and MnS (the predominant forms of AVS), respectively, thus resulting in the reduction of AVS content. The addition of nitrate also caused the oxidation of organic matter (Ripl, 1976; Jenneman et al., 1986). Huang et al. (2010) illustrated that the TOC in the sediment decreased by 10.5% and 9.5% in a reactor −2 , respectively. The results obtained in with 140 and 70 g NO− 3 -N·m 100
(a) a
a
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type 3
Fig. 4. Comparison of (a): AVS and (b): TOC of the sediments on the 28th day after different treatments (n = 3).
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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type 3
type 2
type 1
control
Pres H2SO4-P BD-P NaOH-P NH4Cl-P
start
0
500
1000
1500
P (mg·kg -1)
2000
2500
Fig. 5. Comparison of the distributions of different P binding forms in the sediments on the 28th day after different treatments (n = 3).
the present work (Fig. 4b) are consistent with the results of Huang et al. (2010). The contained N and P compounds were apparently released at a similar ratio by the degradation of organic matter (Sinke et al., 1990), thus contributing to higher N and P concentrations in the water column. During the later stage of experiment, the colour of the upper sediment gradually turned from dark to a yellowish orange. This phenomenon indicated the presence of oxide–hydroxides of Fe(III), which replaced the reduced forms of Fe(II) (Yamada et al., 2012). The formation of calcium phosphate in the sediment caused by the addition of calcium nitrate could also contribute to the change in the surface sediment layer. The colour of deep layer sediment could change as the reaction progresses, and this phenomenon was also described in the study of Yamada et al. (2012). With the variation in sediment colour, the release of sulphur-containing odour is expected to be reduced significantly by the addition of calcium nitrate. 4.2.2. Phosphorus The results demonstrated that the addition of calcium nitrate caused variation in the P fraction of the sediment and led to a decline in the PO34 −-P concentration of the water column. The recycling of P from the bottom sediment to the water column can be affected by many factors, including pH, ORP, mineralisation and microbial activities (Boström et al., 1988; Burley et al., 2001). Different factors can be dominant in various places. The P concentration in the water column was directly attributed to the combination of mineralisation of organically bound P and the desorption/dissolution of redox-sensitively bound P, both of which are the most important P mobilisation processes in the sediments (Schauser et al., 2006). Mineralisation is a temperature-controlled process, whereas the desorption/dissolution of redox-sensitively bound P is mainly redox controlled and is bound to iron oxyhydroxides (Schauser et al., 2006). Thus, the variation in P with the addition of nitrate in this study was mainly caused by the latter. The release potential of P in the sediment can be estimated using profiles of the labile P fractions NH4Cl-P, BD-P and NaOH-P. In Lakes Volvi and Koronia in Greece, Ca-P (HCl-P) had the largest contribution to sedimentary inorganic P loads (Kaiserli et al., 2002), whereas iron compounds and organic matter seemed to play significant roles in regulating the NaOH-P budget. Comparatively, the P fraction in the Shajing river sediment was mostly NaOH-P and H2SO4-P in our study. The PO3− 4 -P concentration of the water column in the control exhibited an increasing trend. The reason could be the release of P from the mineralisation of organically bound P, which is more dominant than the desorption or sedimentation of P in the water column. However, the input of calcium nitrate increased the ORP of the sediment and led to the formation of Fe(III) and SO24 −. The added calcium and Fe(III)
7
released in the sediment contributed not only to the increase in H2SO4-P and BD-P in the sediment and, but also to the decrease in NH4Cl-P (labile fraction P), thus leading to a low P release rate from the sediment. Consequently, the calcium nitrate added to the sediment resulted in the decreased PO3− 4 -P concentration of the water column. In our study, the fraction of NaOH-P increased because the reoxidation of Fe(II) occurred in the oxygenated water with subsequent co-precipitation with Fe(OH)3 and adsorption of PO34 − with Fe(III) (Ripl, 1976; Cooke et al., 1993; Gonsiorczyk et al., 2001; Janke et al., 2011), whereas the fraction of H2SO4-P increased due to the formation of Ca–P compounds. The stability of Fe–P compounds was strongly dependent on the changes in the redox state. 4.2.3. Nitrogen Compared to the control, the injection of calcium nitrate into the river sediment significantly increased the NO− 3 -N concentration in the water column and the interstitial NO− 3 -N in the sediment. This increase may become a new potential risk in the aquatic environment. The results from the current study are consistent with those of previous works (Feibicke, 1997; Yamada et al., 2012). Yamada et al. (2012) found that the NO− 3 -N concentrations in the interstitial water and water column with the treatments were significantly higher than those in the control due to the solubility and the continuous release of added nitrate. The researchers (Yamada et al., 2012) found that the −1 on the 85th day. A rehighest NO− 3 -N concentration was 285 mg·L − duction of approximately 87% in the NO3 -N concentration was observed after this period towards the end of the experiments. The added nitrate was denitrified after several weeks, and a low level of nitrate was observed. In the present work, the highest NO− 3 -N concentration in the water column (11.9 mg·L−1) was lower than that stated in previous works (Janke et al., 2011; Yamada et al., 2012). This could be due to the difference in the dosage of calcium nitrate. Furthermore, the results from current study did not present a submit point and a decreasing trend of NO− 3 -N concentration like the study of Yamada et al. (2012). Such outcome could be attributed to the shorter incubation time (28 days) than the 145 days in the above mentioned research (Yamada et al., 2012). The interstitial NH+ 4 -N concentration in the sediment decreased at the end of the experiment irrespective of the treatments, whereas it was lower than that in the control on the 28th day (Fig. 3a). This phenomenon could explain the result that the NH+ 4 -N concentration of the water column increased during the initial stage (10 days). The formation of NH+ 4 -N could be linked to the dissimilatory reduction in the anoxic and sulphide-containing environments (Brunet and Garcia-Gil, 1996). Similar results were found in Yamada et al. (2012), indicating that the NH+ 4 -N concentrations of water samples with calcium nitrate were significantly higher than that of the control, and that the highest NH+ 4 -N concentration occurred on the 5th day after the addition of calcium nitrate. In the current work, the NH+ 4 -N concentration of the water column decreased after 10 days. This finding could also be correlated with the continuous ammonia volatilisation loss and the transformation + of NH+ 4 -N to N2. Furthermore, NH4 -N could be oxidized to nitrate by heterotrophic microorganisms because of the improvement of redox potential, resulting in the reduction of NH+ 4 -N of the water column after the 10th day. According to the results of Yamada et al. (2012), the addition of calcium nitrate significantly increased the TN concentration on the 0th day. However, the TN differences between the control and treatment microcosms were not significant on the 145th day. In the current study, the addition of calcium nitrate increased the TN concentration of the water column on the 28th day. Gas bubbles were also observed during the experiments. The result obtained by Foy (Foy, 1986) indicated that denitrification to nitrogen gas was the dominant reaction. Gas bubbles indicated the occurrence of an intense denitrification process caused by the addition of nitrate, which could explain the reduction of TN of the water column.
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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5. Conclusions The addition of calcium nitrate into the river sediment improved the ORP of the sediment, thus resulting in the transformation from an anoxic to an oxidizing sediment environment. The AVS content in the sediment was significantly reduced (by 54.9% to 57.1%) due to the content in oxidation of sulphide. Due to this, a high interstitial SO2− 4 concentration in the water column were obthe sediment and SO2− 4 served. The colour of the upper sediment gradually turned to a yellowish orange after the addition of calcium nitrate, which contributed to controlling the sulphur-containing odour. The addition of calcium nitrate also improved the binding capacity of P in the sediment and reduced the release of P from the sediment, leading to a low P concentration of the water column. The addition of calcium nitrate also caused a high NO− 3 -N concentration in the water column and interstitial NO− 3 -N content in the sediment. Water quality improvement had a limited effect on the remediation of sediment with calcium nitrate, which significantly affected only the interstitial nitrate and sulphate concentrations in the sediment. In future, field experiments will be conducted to provide guidelines for the remediation of sediment. Acknowledgement This study was financially supported by PhD Start-up Fund of National Science Foundation of Guangdong Province, China (2016A030310024), China Postdoctoral Science Foundation (2015M581116), Science and Technology Department, Guangdong Province (2015A010106002) and Research Grant from Graduate School at Shenzhen, Tsinghua University (No. JC2015001). We also express our appreciation to Hairong Wen and Shiwen Huang from Shenzhen University for the analyzing work in the lab. References APHA, 1998. Standards Methods for the Examination of Water and Wastewater. 19th ed. American Public Health Association/American Water Works Association/, Washington, DC, USA. Berner, R.A., 1964. Distribution and diagenesis of sulfur in some sediments from the Gulf of California. Mar. Geol. 1 (2), 117–140. Berner, R.A., 1970. Sedimentary pyrite formation. Am. J. Sci. 268 (1), 1–23. Boström, B., Andersen, J.M., Fleischer, S., Jansson, M., 1988. Exchange of phosphorus across the sediment–water interface. Hydrobiologia 48, 229–244. Brunet, R.C., Garcia-Gil, L.J., 1996. Sulfide-induced dissimilatory nitrate reduction to ammonia in anaerobic freshwater sediments. FEMS Microbiol. Ecol. 21, 131–138. Burley, K.L., Prepas, E.E., Chambers, P.A., 2001. Phosphorus release from sediments in hardwater eutrophic lakes: the effects of redox-sensitive and -insensitive chemical treatments. Freshw. Biol. 46, 1061–1074. Caraco, N.F., Cole, J.J., Likens, G.E., 1989. Evidence for sulphate-controlled phosphorus release from sediments of aquatic systems. Nature 341, 316–318. Chen, L., Wang, L., Liu, S., Zhang, X., Hu, J., Tao, Y., 2013. Effect of calcium nitrate on odor and properties of chemistry in sediment of Shenzhen River. J. Harbin Institute Technol. 45 (6), 107–113 In Chinese. Cooke, G.D., Welch, E.B., Martin, A.B., Fulmer, D.G., Hyde, J.B., Schrieve, G.D., 1993. Effectiveness of Al, Ca, and Fe salts for control of internal phosphorus loading in shallow and deep lakes. Hydrobiologia 84, 323–335. Feibicke, M., 1997. Impact of nitrate addition on phosphorous availability in sediment and water column on plankton biomass - experimental field study in the shallow brackish Schlei Fjord (Western Baltic, Germany). Water Air Soil Pollut. 99, 445–456. Foy, R.H., 1986. Suppression of phosphorous release from lake sediments by the addition of nitrate. Water Res. 20 (11), 1345–1351. Fulton III, R.S., Godwin, W.F., Schaus, M.H., 2015. Water quality changes following nutrient loading reduction and biomanipulation in a large shallow subtropical lake, Lake Griffin, Florida, USA. Hydrobiologia 753, 243–263. Funes, A., de Vicente, J., Cruz-Pizarro, L., Alvarez-Manzaneda, I., de Vicente, I., 2016. 2016. Magnetic microparticles as a new tool for lake restoration: a microcosm experiment for evaluating the impact on phosphorus fluxes and sedimentary phosphorus pools. Water Res. 89, 366–374. Gonsiorczyk, T., Casper, P., Koschel, R., 2001. Mechanisms of phosphorus release from the bottom sediment of the oligotrophic Lake Stechlin: importance of the permanently oxic sediment surface. Arch. Hydrobiol. 151 (2), 203–219.
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Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
Contents lists available at ScienceDirect
Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv
Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate Xiaoning Liu a,b, Yi Tao a,b,⁎, Kuiyu Zhou a, Qiqi Zhang a, Guangyao Chen a, Xihui Zhang a,b a b
Key Laboratory of Microorganism Application and Risk Control of Shenzhen, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China Tsinghua-Kangda Research Institute of Environmental Nano-Engineering & Technology, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China
H I G H L I G H T S
G R A P H I C A L
A B S T R A C T
• AVS in the sediments decreased significantly with the addition of calcium nitrate. • Calcium nitrate contributed to the reduction of phosphate in the water column. • Water quality had effect on the interstitial nitrate and sulfate in the sediment.
a r t i c l e
i n f o
Article history: Received 18 July 2016 Received in revised form 17 September 2016 Accepted 17 September 2016 Available online xxxx Editor: Jay Gan Keywords: Calcium nitrate Sediment Water quality Acid volatile sulphide Phosphorus Total organic carbon
a b s t r a c t In situ sediment remediation technique is commonly used to control the release of pollutants from sediment. Addition of calcium nitrate to sediment has been applied to control the release of phosphorus from sediments. In this study, laboratory experiments were conducted to investigate the effect of water quality improvement on the remediation of river sediment with the addition of calcium nitrate. The results demonstrated that the redox-potential of sediments increased from −282 mV to −130 mV after 28 days of calcium nitrate treatment. The acid volatile sulphide in the sediments significantly decreased (by 54.9% to 57.1%), whereas the total organic carbon decreased by 9.7% to 10.2%. However, the difference between these and water quality improvement was not significant. Due to the addition of calcium nitrate, low phosphate concentration in the water column and interstitial phosphate in the sediment were observed, indicating that the calcium nitrate was beneficial to controlling the release of phosphorus from river sediment. The decrease in phosphorus release could be attributed to the fixation of iron–phosphorus and calcium–phosphorus due to the addition of calcium nitrate. The addition of calcium nitrate to sediment caused the oxidation of sulphide to sulphate, hence resulting in high nitrate and sulphate concentrations in the water column, and high interstitial nitrate and sulphate concentrations in the sediment. The results also showed that only the water quality improvement had a significant effect on the interstitial nitrate and sulphate concentrations in the sediment. © 2016 Elsevier B.V. All rights reserved.
1. Introduction ⁎ Corresponding author at: Graduate School at Shenzhen, Tsinghua University, China. E-mail address: [email protected] (Y. Tao).
Rivers have been the receivers of domestic wastewater effluent, rainwater runoff, agricultural runoff and industrial wastewater for
http://dx.doi.org/10.1016/j.scitotenv.2016.09.149 0048-9697/© 2016 Elsevier B.V. All rights reserved.
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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years, and this has seriously damaged the river water quality (Scholes et al., 2008). Due to rapid increases in economy and population in China, the quantity of pollutants discharged into the rivers has also increased. The continuing development of pollution in water bodies causes them to become turbid and odourous after their carrying capacity of these pollutants has been exceeded. Water pollution mostly occurs in water bodies, which either have a low flow rate or are static. Therefore, water pollution of urban rivers is becoming one of the greatest threats to the aquatic ecosystem in China. Large amounts of organic pollutants from wastewater are accumulated in river sediment, while excessive degradation of the organic matter consumes much dissolved oxygen (DO) resulting in hypoxia of the sediment. Low DO causes a low redox potential [oxidation reduction potential (ORP)], which leads to microbial reduction of sulphate (SO2− 4 ) to hydrogen sulphide (H2S) in the sediment (Berner, 1970). A decrease in the ORP of sediment can also cause the reduction of Fe and Mn, which results in the release of phosphorus (P) adsorbed on Fe and Mn oxide–hydroxide particles (Caraco et al., 1989; Schauser et al., 2006). Subsequently, Fe(II) in the water column and sediment reacts with H2S to form iron sulphide (FeS) because of its low-solubility product and gets adsorbed on the organic particles, causing the water and sediment layer to turn black-coloured. Furthermore, toxic substances, such as ammonia, H2S and other odourous gaseous sulphur compounds, are produced as a result of anaerobic respiration, which leads to odourous phenomenon along the rivers. Water quality can be improved with the mitigation of external pollution sources and advanced treatment of wastewater. However, sediments can continuously release pollutants to the water column, thus delaying the recovery of eutrophic environments (Søndergaard et al., 2007). Previous research indicated that eutrophic ecosystems could be recovered using in situ technologies, such as addition of nitrate, Al, Fe and Ca to the sediment (Ripl, 1976; Foy, 1986; Cooke et al., 1993; Kleeberg et al., 2000; Xu et al., 2008; Yin and Kong, 2015). The addition of these compounds to sediments can improve the oxidation conditions and accelerate the denitrification activity (Shimizu and Nakano, 2009). Ripl (1976) found that the biochemical oxidation could be achieved for remediation purposes by injecting calcium nitrate into the sediment to control P release from the sediment, which would limit the growth of algae in the water body. Apart from controlling the P release from sediment, nitrate is also used to abate the odours caused by sulphate-reducing bacteria in wastewater (Jenneman et al., 1986; Jefferson et al., 2002; Jiang et al., 2009). When nitrate is metabolised, a portion of the organic matter is used to to S2− (Jenneman et al., 1986). The acid volatile sulphide reduce SO2− 4 (AVS) concept is defined as the sedimentary S, which is extracted by 1 mol·L−1 HCl solution, including porewater sulphides and metastable FeS (Berner, 1964). The AVS content in the sediment is strongly correlated with the colour intensity of the sediment, which could be used to assess the pollution of organically enriched sediment (Wilson and Vopel, 2012). Sulphur-containing odourous compounds also originate from AVS in the sediment, and to some extent, AVS can be considered as the source of sulphur-containing odour and potential production of H2S. Thus, reducing AVS content in the sediment and controlling the release of sulphur-related odour to the water column are meaningful and urgent research tasks. P loading is considered to have a key role in the formation of algal blooms and eutrophication (Yin et al., 2013). Since P release from anaerobic sediments usually makes up the major fraction of total P (TP) load in lakes or rivers, therefore the reduction of an external P load by controlling the influent water quality may be insufficient (Cooke et al., 1993). Consequently, a remediation planning to control the P sources should be a priority. Many amendment technologies, such as modified zeolite (Yang et al., 2014), magnetic microparticles (Funes et al., 2016) and thermally modified calcium-rich attapulgite (Yin et al., 2013; Yin and Kong, 2015) have been processed to immobilize both the soluble reactive P in porewater and mobile P in sediment.
Upgradation of wastewater treatment plants and the mitigation of external pollution sources have been considered effective in improving the lake and river water quality (Jeppesen et al., 2007; Fulton et al., 2015). However, both the improvement and remediation of polluted sediment are far slower than that of the water column. Meanwhile, sediment would continue to be an important internal source of nutrients to the water column for a long time, which affects the water quality of overlying water due to a larger nutrient concentration in the sediment porewater than in the water column (Liu et al., 2014). In the present study, the effect of water quality improvement on sediment remediation has been investigated, which is related to the variation in AVS due to the addition of calcium nitrate. 2. Material and methods 2.1. Water and sediment sampling The sediment and water samples used in this study were obtained from the Shajing River (22°46′3″, 113°49′52″) in Shenzhen City, Guangdong, China. This river is a tributary of the Maozhou River, which flows into the Pearl River Estuary. In the present study, the water and sediment samples were obtained from upstream, midstream and downstream of the Shajing River, and were mixed to present the characteristics of water and sediment. The river flows slowly at the depth of 1.5–2.0 m, and receives pollutants of rainfall runoff and occasional industrial and urban wastewaters from the riverside. The water samples were obtained from the middle zone of the river using a glass water sampler, while the sediment samples were collected from the top 20 cm layer of three sites using a piston column sediment sampler (purchased from Nanjing Institute of Geography and Limnology, Chinese Academy of Sciences, China). The samples were homogenized and sieved to remove detritus, plant residues and plastic products. Then, the collected samples were kept cool in a box containing ice cake. These were transferred to the laboratory and stored at a temperature of approximately 4 °C until further use. The water quality and sediment's characteristics of the sampling station are summarized in Tables 1 and 2. 2.2. Experimental design The experiments were conducted using 1 L wide-mouth jars having screw top lids. Before the experiments, the jars were placed in a sterilizing pot at 121 °C for 30 min. The sediment (250 g) and water column (800 mL) were placed in the jars, and subsequent treatment was performed in triplicate. The heights of sediment and overlying water were 5 and 10 cm, respectively. Four treatments were conducted. The first treatment did not receive calcium nitrate (control treatment). The sediments in the other three treatments were given the same dosage of calcium nitrate at three different water column compositions as shown in Table 3. The water column was replaced with pure water, which represented the treated water. Based upon the study of Chen et al. (2013), the application dosage was 1 g of calcium nitrate per 1 kg of wet sediment. Nitrogen gas was injected into the jars when the calcium nitrate was mixed with the sediment to avoid sediment oxidation. The jars were kept closed during the experiments, which were performed at 25 °C in an incubator. The water samples were withdrawn on 0th, 1st, 3rd, 6th, 10th, 15th, 21st and 28th day of the experiment, while those of sediment were withdrawn on 0th and 28th day of the Table 1 Characteristics of the Shajing River water investigated in this study. Unit
pH
DO mg·L
Mean Std
7.26 0.05
0.26 0.04
Eh −1
NH+ 4 -N
mV
mg·L
−68.6 12.2
14.6 1.2
−1
NO− 3 -N mg·L 0.10 0.02
−1
TN mg·L 18.2 0.86
−1
PO3− 4 -P
COD
mg·L−1
mg·L−1
6.2 1.2
84.5 15.0
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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Table 2 Characteristics of the sediments obtained from the Shajing River. Sediment
pH
Unit Mean Std
7.25 0.12
Eh
Water content
Total organic carbon
Acid volatile sulphide
mV
%
mg·g−1
mg S·kg−1
−282 35
78.2 3.2
90.4 3.4
5136 102
experiment. The measured values in the sediment before the experiment (0th day) are defined as “start” values. Nitrogen gas was injected into the flask to drive out oxygen prior to shaking. The diluted slurry was centrifuged (7000 rpm·min− 1), while the supernatant was − filtrated through 0.45 μm Millipore filters for measuring NH+ 4 -N, NO3 2− -P and SO . N, PO3− 4 4 2.3. Measurement method and statistical analysis The water samples were filtered through 0.45 μm Millipore filters before chemical analysis to determine the dissolved nutrients. The pH and DO concentrations of the water column were measured using a Hach HQ40d Dual-input Multi-parameter Meter. According to the standard methods described by the American Public Health Association − 3− 2− (1998), the NH+ 4 -N, NO3 -N, total nitrogen (TN), PO4 -P and SO4 were analyzed by colorimetric analysis by using a spectrophotometer. The pH and ORP of the sediment were measured with a Hach HQ40d Dual-input Multi-parameter Meter at a depth of 3 cm from the surface layer on 0th and 28th day of the experiment. The dry weight was measured by drying the sediment at 105 °C for 24 h, while the water content was calculated based upon the weight loss. The interstitial ions in the sediments were extracted by diluting the sediment sample in 1:4 ratios using pure water. The diluted sample was stirred for 4 h at 200 rpm·min−1. Nitrogen gas was injected into a closed conical flask to maintain the hypoxia condition before shaking it in the shaker. The AVS in the sediment was analyzed using the method described by Hsieh and Shieh (1997). The TOC in the sediment was measured using potassium dichromate titration method (Schumacher, 2002). The fractional composition of TP (different P binding forms) in the sediment was examined using the fractionation scheme of Psenner et al. (1984). This sequential extraction method divides the TP of sediment into pools as shown in Table 4. The amount of P in each fraction is determined as TP described above. The difference between TP and the sum of the extracted pools is the residual P (Pres), which consists mainly of organic bound P. The average values and standard deviations of three replicate samples were calculated. Statistical analysis (two-way ANOVA) was performed to evaluate the significant differences among the means in the sediment by using SPSS Statistics 19.0. 3. Results 3.1. Parameters of the water column The pH value of the water column for all samples exhibited a transient increasing trend before it began to decline consistently (Fig. 1a). The pH values of the treatments with added calcium nitrate were slightly higher than that in the control on the 28th day. The mean pH value in Table 3 The experimental design and water column composition for the four treatments. Treatments
Experimental design
Water column composition
Control Type 1 Type 2
Sediment Sediment and Ca(NO3)2 Sediment and Ca(NO3)2
Type 3
Sediment and Ca(NO3)2
800 mL of water samples 800 mL of water samples 720 mL of water samples and 80 mL of pure water 600 mL of water samples and 200 mL of pure water
the control was 6.82, while those in types 1, 2 and 3 experiments were 7.08, 7.08 and 7.06, respectively. Neither the addition of calcium nitrate to sediment nor water quality improvement had any obvious effect on the pH of the water column. The DO values of the water column in the control showed no significant variation during the entire experiment. However, those of the treatments exhibited a sharply increasing trend during the first several days. The DO values in types 1 and 2 experiments reached the highest values (2.03 and 2.64 mg·L− 1, respectively) on the 3rd day, while type 3 treatment (2.15 mg·L−1) reached the highest value on the 6th day. These values decreased to a lower value on the 10th day, ending up even lower than 0.5 mg·L−1 at the end of the experiment (Fig. 1b). The addition of calcium nitrate to sediment affected only the DO value to some extent during the first several days, while the water quality improvement had no obvious effect on the DO value. The NO− 3 -N concentration in the control showed an increasing trend, whereas that of the calcium nitrate treatments first increased sharply for a short time period (1 day), decreased for 10 days and then increased again with time (Fig. 2a). The concentrations increased to 8.8, 10.4 and 11.9 mg·L−1 on the 28th day in types 1, 2 and 3 experiments, respectively. Overall, the addition of calcium nitrate increased the NO− 3 -N concentration in the water column, and the NO− 3 -N concentration in type 2 experiment exhibited the highest value (11.9 mg·L− 1). The NH+ 4 -N concentration in the water column increased sharply during the initial stage of the experiment, reaching a maximum of nearly 40 mg·L−1 on the 10th day. Then, the concentration began to decrease consistently (Fig. 2b). The NH+ 4 -N concentrations in the calcium nitrate treatments were slightly higher than that in the control on the 28th day. The TN concentration exhibited a similar trend. It reached the maximum value of 44.0, 47.6, 46.1 and 46.1 mg·L−1 on the 10th day in control, types 1, 2 and 3 experiments. Afterwards, the concentration began to decrease consistently (Fig. 2c). The TN concentration in the water column in the control was slightly lower than those in the treatments. However, all of these reached the maximum values on the 10th day. The PO34 −-P concentration in the water column in the control decreased during the initial period (1st day) and then increased consistently with time (Fig. 2d), reaching a maximum value of 2.4 mg·L−1 on the 10th day. The concentration subsequently exhibited a slowly declining trend and decreased to 2.0 mg·L−1 on the 28th day. On the other hand, the PO3− 4 -P concentrations exhibited a decreasing trend in the calcium nitrate treatments, and the values on the 28th day were found to be 1.2, 0.7 and 0.7 mg·L−1 in types 1, 2 and 3 experiments, respectively. The SO24 − concentration in the water column decreased slowly from 98.5 mg·L−1 to 79.3 mg·L−1 in the control, whereas it exhibited an increasing trend in the calcium nitrate treatments (Fig. 2e), reaching 138.1, 150.8 and 119.2 mg·L−1 on the 28th day in types 1, 2 and 3 experiments, respectively.
Table 4 Sequential extraction methods of different P binding forms in the sediment. Mark
Extraction/Time
Expected P binding forms
NH4Cl-P NaOH-P BD-P H2SO4-P Pres
1.0 M NH4Cl/2 h 0.1 M NaOH/16 h 0.3 M Na2S2O4/NaHCO3/1 h 0.25 M H2SO4/2 h
Pore water P, Loosely adsorbed to surfaces Inorganic P bound to Al Redox-sensitive P Ca-binding P Organic and other refractory P
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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X. Liu et al. / Science of the Total Environment xxx (2016) xxx–xxx 7.6
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3.2. Interstitial ions in the sediment The variations in the interstitial ions in the sediment are presented in Fig. 3. The interstitial NO− 3 -N concentration in the sediment in the control increased from 6.1 mg·kg−1 to 7.1 mg·kg− 1. The interstitial NO− 3 -N concentrations increased significantly to values of 22.9, 18.1 and 19.0 mg·kg−1 in types 1, 2 and 3 experiments, respectively (Fig. 3a). The interstitial NO− 3 -N concentrations in the sediment in types 1, 2 and 3 experiments were significantly different from that in the control. Type 1 experiment had the greatest effect on the interstitial NO− 3 N concentration in the sediment, which was also significantly different from those in types 2 and 3 experiments. The interstitial NH+ 4 -N concentration in the sediment of the control decreased from 646.7 mg·kg−1 to 472.7 mg·kg−1, whereas the corresponding values in types 1, 2 and 3 experiments on the 28th day were found to be 406.7, 385.3 and 393.0 mg·kg−1, respectively. The interstitial NH+ 4 -N concentration in the sediment in the control was higher than that in types 1, 2 and 3 experiments (Fig. 3b), and was found to be significantly different from those in types 2 and 3 experiments. The interstitial PO3− 4 -P concentration in the sediment of the control decreased from 2.44 mg·kg− 1 to 2.00 mg·kg−1, whereas the corresponding values in types 1, 2 and 3 experiments were found to be 0.36, 0.27 and 0.41 mg·kg−1, respectively. The interstitial PO3− 4 -P concentration in the sediment in types 1, 2 and 3 experiments were found to be significantly different from that in the control. The intersticoncentration in the sediment of the control increased from tial SO2− 4 787 mg·kg−1 to 813 mg·kg−1, whereas the corresponding values for types 1, 2 and 3 experiments were found to be 1012, 1205 and 1201 concentrations in the sedmg·kg−1, respectively. The interstitial SO2− 4 iments in types 2 and 3 experiments were significantly different from that of the control; however the value was not significantly different from that in type 1 experiment.
3.3. Variations in AVS and TOC in the sediment The AVS content in the control was slightly lower than that in the “start” (before the experiment). Compared with that in the control, the AVS content in the sediments of treatments decreased significantly with the addition of calcium nitrate. The corresponding values in types 1, 2 and 3 experiments decreased from 5136 mg S·kg−1 to 2205, 2313 and 2223 mg S·kg−1, respectively (Fig. 4a). However, no significant difference was observed among the values of three treatments with different water qualities. The variation in TOC of the sediments is presented in Fig. 4b. No obvious difference was observed in the TOC content of the sediment in the control (the value decreased from 90.4 mg·kg− 1 to 88.1 mg·kg− 1). However, the TOC contents significantly decreased
from 90.4 mg·kg−1 to 81.2, 81.6 and 81.2 mg·kg−1 in types 1, 2 and 3 experiments, respectively. Similarly, water quality had no significant effect on the TOC content of the sediment. 3.4. P fraction in the sediment The distributions of different P binding forms after 28 days of the addition of calcium nitrate are presented in Fig. 5. The fraction of P binding forms changed, however water quality had no significant effect on them. No obvious difference was observed in the NH4Cl-P in the sediment of the control. On the other hand, NH4Cl-P decreased from 62.95 mg·kg−1 to 5.32, 5.55 and 5.19 mg·kg−1 for the types 1, 2 and 3 experiments, respectively. In the original sediment, P was mostly NaOH-P and H2SO4-P, while BD-P and NH4Cl-P accounted for only a small fraction. After adding calcium nitrate, the amount of NH4Cl-P in the sediment decreased, whereas the amounts of NaOH-P, H2SO4-P and BD-P increased. Furthermore, Pres exhibited a slightly decreasing trend in all treatments. 4. Discussion 4.1. Effect of water quality Water quality of the river had a certain effect on the interstitial NO− 3 N concentration in the sediment and on the NO− 3 -N concentration in the water column. Unlike in type 1 experiment, higher NO− 3 -N concentration of water column was found in types 2 and 3 experiments, thus indicating their improved water quality compared with that in type 1 experiment. However, water quality had insignificant effect on the in+ terstitial NH+ 4 -N concentration in the sediment and on the NH4 -N and TN concentrations in the water column. Furthermore, water quality had no significant effect on the AVS and TOC in the sediment (Fig. 4). The PO3− 4 -P concentrations in types 2 and 3 experiments on the 28th day were lower than that in type 1 experiment (Fig. 2d). Nevertheless, water quality had no significant effect on the variation of P fraction in the sediment. Similarly, with the improvement of water quality, a lower interstitial PO3− 4 -P in the sediment in types 2 and 3 experiments were observed than that in type 1 experiment, though there was no significant difference between the values. Thus, the variation in water quality had no obvious effect on both the interstitial PO3− 4 -P in the sediment and PO3− 4 -P concentration of the water column. Sediment nutrient concentrations are generally weakly correlated with water quality of the lake (Trolle et al., 2009). This means that the water quality has no obvious effect on the variation in sediment pollutants. However, improvement of water quality theoretically increases the diffusive flux at the sediment–water interface, thus leading to a higher release rate of pollutants from the sediment. The results of variation in several major pollutants of water column did now show such
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
X. Liu et al. / Science of the Total Environment xxx (2016) xxx–xxx 50
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trend (Fig. 2). Although the addition of calcium nitrate had an obvious effect on the variation in ions in the water column, water quality improvement had no obvious influence on the remediation of river sediment due to the addition of calcium nitrate, particularly on the variation in the AVS content in the sediment. Given the short incubation time (28 days) in this study, a longer time is required to assess the function of water quality improvement in remediating the river sediment. 4.2. Effect of calcium nitrate 4.2.1. Sulphur The AVS content in the sediment significantly decreased by 54.9% to 57.1%, while the TOC in the sediment decreased by 9.7% to 10.2% (Fig. 4). Janke et al. (2011) also found that the AVS content in the sediment decreased by 99% after 135 days. Yamada et al. (2012) found that the AVS content in the sediment, which was treated with calcium nitrate,
decreased by 99%. However, the AVS content in the sediment treated with calcium nitrate, was reduced by only 54.9% to 57.1% (Fig. 4), and exhibited a lower reduction rate than that indicated in the literature (Janke et al., 2011; Yamada et al., 2012). Firstly, the reason might be the different in dosages of calcium nitrate as the denitrification activity is proportional to the nitrogen concentration (Wang et al., 2007). Secondly, the loss of nitrate caused by the high water solubility of calcium nitrate in the sediment water could not be directly measured. The nitrate loss may be estimated by both the increase in NO− 3 -N concentration in the water column and interstitial NO− 3 -N in the sediment. Lastly, the current study used a shorter incubation time (28 days) than that of the Yamada et al. (2012), who used a 145 days incubation time. The AVS content in the sediment is related to seasonal and depth variations of TOC, sediment redox potential (Eh) and sulphate concentration in the sediment cores and porewater (Yin and Fan, 2011). The
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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reduction in AVS content in the sediment was mainly attributed to two reasons. Firstly, the levels of sulphate-reducing bacteria decreased during the prolonged exposure to an oxidizing environment and high N2O concentration (Mahmood et al., 2007). Due to these, the microbial reduction of SO24 − to H2S was inhibited. The ORP of the sediment in types 1, 2 and 3 experiments in this study increased from − 282 mV to −130, −127.4 and −128.2 mV, respectively, thus resulting in the inhibition of sulphate-reducing bacteria. The decrease in the ORP of the sediment was caused by the build-up of produced N2O and NO, while the reduction in the amount of organic matter and the production of N2 during denitrification resulted in an oxidizing environment (Jenneman et al., 1986). Secondly, nitrate acted as an electron acceptor, whereas the sulphide in the sediment was oxidized to SO2− 4 in the presence of nitrate (Ripl, 1976; Jenneman et al., 1986). Jiang et al. (2009) assumed that sulphide was oxidized in two steps under anoxic conditions
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after the addition of nitrate. The first step was the oxidation of sulphide to S, while the second step was further oxidation of elemental sulphur to SO2− 4 . The two steps occurred sequentially with the second step occurring primarily after sulphide depletion, whereas the first step was considerably faster than the second step. This inference could be supported by both the increase in SO2− 4 concentration in the water column in the sediment (Fig. 3d). Janke et al. (Fig. 2e) and the interstitial SO2− 4 (2011) also stated that the oxidation of sulphide would increase both concentration and the availability of Fe and Mn, which were the SO2− 4 previously complexed as FeS and MnS (the predominant forms of AVS), respectively, thus resulting in the reduction of AVS content. The addition of nitrate also caused the oxidation of organic matter (Ripl, 1976; Jenneman et al., 1986). Huang et al. (2010) illustrated that the TOC in the sediment decreased by 10.5% and 9.5% in a reactor −2 , respectively. The results obtained in with 140 and 70 g NO− 3 -N·m 100
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Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
X. Liu et al. / Science of the Total Environment xxx (2016) xxx–xxx
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the present work (Fig. 4b) are consistent with the results of Huang et al. (2010). The contained N and P compounds were apparently released at a similar ratio by the degradation of organic matter (Sinke et al., 1990), thus contributing to higher N and P concentrations in the water column. During the later stage of experiment, the colour of the upper sediment gradually turned from dark to a yellowish orange. This phenomenon indicated the presence of oxide–hydroxides of Fe(III), which replaced the reduced forms of Fe(II) (Yamada et al., 2012). The formation of calcium phosphate in the sediment caused by the addition of calcium nitrate could also contribute to the change in the surface sediment layer. The colour of deep layer sediment could change as the reaction progresses, and this phenomenon was also described in the study of Yamada et al. (2012). With the variation in sediment colour, the release of sulphur-containing odour is expected to be reduced significantly by the addition of calcium nitrate. 4.2.2. Phosphorus The results demonstrated that the addition of calcium nitrate caused variation in the P fraction of the sediment and led to a decline in the PO34 −-P concentration of the water column. The recycling of P from the bottom sediment to the water column can be affected by many factors, including pH, ORP, mineralisation and microbial activities (Boström et al., 1988; Burley et al., 2001). Different factors can be dominant in various places. The P concentration in the water column was directly attributed to the combination of mineralisation of organically bound P and the desorption/dissolution of redox-sensitively bound P, both of which are the most important P mobilisation processes in the sediments (Schauser et al., 2006). Mineralisation is a temperature-controlled process, whereas the desorption/dissolution of redox-sensitively bound P is mainly redox controlled and is bound to iron oxyhydroxides (Schauser et al., 2006). Thus, the variation in P with the addition of nitrate in this study was mainly caused by the latter. The release potential of P in the sediment can be estimated using profiles of the labile P fractions NH4Cl-P, BD-P and NaOH-P. In Lakes Volvi and Koronia in Greece, Ca-P (HCl-P) had the largest contribution to sedimentary inorganic P loads (Kaiserli et al., 2002), whereas iron compounds and organic matter seemed to play significant roles in regulating the NaOH-P budget. Comparatively, the P fraction in the Shajing river sediment was mostly NaOH-P and H2SO4-P in our study. The PO3− 4 -P concentration of the water column in the control exhibited an increasing trend. The reason could be the release of P from the mineralisation of organically bound P, which is more dominant than the desorption or sedimentation of P in the water column. However, the input of calcium nitrate increased the ORP of the sediment and led to the formation of Fe(III) and SO24 −. The added calcium and Fe(III)
7
released in the sediment contributed not only to the increase in H2SO4-P and BD-P in the sediment and, but also to the decrease in NH4Cl-P (labile fraction P), thus leading to a low P release rate from the sediment. Consequently, the calcium nitrate added to the sediment resulted in the decreased PO3− 4 -P concentration of the water column. In our study, the fraction of NaOH-P increased because the reoxidation of Fe(II) occurred in the oxygenated water with subsequent co-precipitation with Fe(OH)3 and adsorption of PO34 − with Fe(III) (Ripl, 1976; Cooke et al., 1993; Gonsiorczyk et al., 2001; Janke et al., 2011), whereas the fraction of H2SO4-P increased due to the formation of Ca–P compounds. The stability of Fe–P compounds was strongly dependent on the changes in the redox state. 4.2.3. Nitrogen Compared to the control, the injection of calcium nitrate into the river sediment significantly increased the NO− 3 -N concentration in the water column and the interstitial NO− 3 -N in the sediment. This increase may become a new potential risk in the aquatic environment. The results from the current study are consistent with those of previous works (Feibicke, 1997; Yamada et al., 2012). Yamada et al. (2012) found that the NO− 3 -N concentrations in the interstitial water and water column with the treatments were significantly higher than those in the control due to the solubility and the continuous release of added nitrate. The researchers (Yamada et al., 2012) found that the −1 on the 85th day. A rehighest NO− 3 -N concentration was 285 mg·L − duction of approximately 87% in the NO3 -N concentration was observed after this period towards the end of the experiments. The added nitrate was denitrified after several weeks, and a low level of nitrate was observed. In the present work, the highest NO− 3 -N concentration in the water column (11.9 mg·L−1) was lower than that stated in previous works (Janke et al., 2011; Yamada et al., 2012). This could be due to the difference in the dosage of calcium nitrate. Furthermore, the results from current study did not present a submit point and a decreasing trend of NO− 3 -N concentration like the study of Yamada et al. (2012). Such outcome could be attributed to the shorter incubation time (28 days) than the 145 days in the above mentioned research (Yamada et al., 2012). The interstitial NH+ 4 -N concentration in the sediment decreased at the end of the experiment irrespective of the treatments, whereas it was lower than that in the control on the 28th day (Fig. 3a). This phenomenon could explain the result that the NH+ 4 -N concentration of the water column increased during the initial stage (10 days). The formation of NH+ 4 -N could be linked to the dissimilatory reduction in the anoxic and sulphide-containing environments (Brunet and Garcia-Gil, 1996). Similar results were found in Yamada et al. (2012), indicating that the NH+ 4 -N concentrations of water samples with calcium nitrate were significantly higher than that of the control, and that the highest NH+ 4 -N concentration occurred on the 5th day after the addition of calcium nitrate. In the current work, the NH+ 4 -N concentration of the water column decreased after 10 days. This finding could also be correlated with the continuous ammonia volatilisation loss and the transformation + of NH+ 4 -N to N2. Furthermore, NH4 -N could be oxidized to nitrate by heterotrophic microorganisms because of the improvement of redox potential, resulting in the reduction of NH+ 4 -N of the water column after the 10th day. According to the results of Yamada et al. (2012), the addition of calcium nitrate significantly increased the TN concentration on the 0th day. However, the TN differences between the control and treatment microcosms were not significant on the 145th day. In the current study, the addition of calcium nitrate increased the TN concentration of the water column on the 28th day. Gas bubbles were also observed during the experiments. The result obtained by Foy (Foy, 1986) indicated that denitrification to nitrogen gas was the dominant reaction. Gas bubbles indicated the occurrence of an intense denitrification process caused by the addition of nitrate, which could explain the reduction of TN of the water column.
Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149
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5. Conclusions The addition of calcium nitrate into the river sediment improved the ORP of the sediment, thus resulting in the transformation from an anoxic to an oxidizing sediment environment. The AVS content in the sediment was significantly reduced (by 54.9% to 57.1%) due to the content in oxidation of sulphide. Due to this, a high interstitial SO2− 4 concentration in the water column were obthe sediment and SO2− 4 served. The colour of the upper sediment gradually turned to a yellowish orange after the addition of calcium nitrate, which contributed to controlling the sulphur-containing odour. The addition of calcium nitrate also improved the binding capacity of P in the sediment and reduced the release of P from the sediment, leading to a low P concentration of the water column. The addition of calcium nitrate also caused a high NO− 3 -N concentration in the water column and interstitial NO− 3 -N content in the sediment. Water quality improvement had a limited effect on the remediation of sediment with calcium nitrate, which significantly affected only the interstitial nitrate and sulphate concentrations in the sediment. In future, field experiments will be conducted to provide guidelines for the remediation of sediment. Acknowledgement This study was financially supported by PhD Start-up Fund of National Science Foundation of Guangdong Province, China (2016A030310024), China Postdoctoral Science Foundation (2015M581116), Science and Technology Department, Guangdong Province (2015A010106002) and Research Grant from Graduate School at Shenzhen, Tsinghua University (No. JC2015001). We also express our appreciation to Hairong Wen and Shiwen Huang from Shenzhen University for the analyzing work in the lab. References APHA, 1998. Standards Methods for the Examination of Water and Wastewater. 19th ed. American Public Health Association/American Water Works Association/, Washington, DC, USA. Berner, R.A., 1964. Distribution and diagenesis of sulfur in some sediments from the Gulf of California. Mar. Geol. 1 (2), 117–140. Berner, R.A., 1970. Sedimentary pyrite formation. Am. J. Sci. 268 (1), 1–23. Boström, B., Andersen, J.M., Fleischer, S., Jansson, M., 1988. Exchange of phosphorus across the sediment–water interface. Hydrobiologia 48, 229–244. Brunet, R.C., Garcia-Gil, L.J., 1996. Sulfide-induced dissimilatory nitrate reduction to ammonia in anaerobic freshwater sediments. FEMS Microbiol. Ecol. 21, 131–138. Burley, K.L., Prepas, E.E., Chambers, P.A., 2001. Phosphorus release from sediments in hardwater eutrophic lakes: the effects of redox-sensitive and -insensitive chemical treatments. Freshw. Biol. 46, 1061–1074. Caraco, N.F., Cole, J.J., Likens, G.E., 1989. Evidence for sulphate-controlled phosphorus release from sediments of aquatic systems. Nature 341, 316–318. Chen, L., Wang, L., Liu, S., Zhang, X., Hu, J., Tao, Y., 2013. Effect of calcium nitrate on odor and properties of chemistry in sediment of Shenzhen River. J. Harbin Institute Technol. 45 (6), 107–113 In Chinese. Cooke, G.D., Welch, E.B., Martin, A.B., Fulmer, D.G., Hyde, J.B., Schrieve, G.D., 1993. Effectiveness of Al, Ca, and Fe salts for control of internal phosphorus loading in shallow and deep lakes. Hydrobiologia 84, 323–335. Feibicke, M., 1997. Impact of nitrate addition on phosphorous availability in sediment and water column on plankton biomass - experimental field study in the shallow brackish Schlei Fjord (Western Baltic, Germany). Water Air Soil Pollut. 99, 445–456. Foy, R.H., 1986. Suppression of phosphorous release from lake sediments by the addition of nitrate. Water Res. 20 (11), 1345–1351. Fulton III, R.S., Godwin, W.F., Schaus, M.H., 2015. Water quality changes following nutrient loading reduction and biomanipulation in a large shallow subtropical lake, Lake Griffin, Florida, USA. Hydrobiologia 753, 243–263. Funes, A., de Vicente, J., Cruz-Pizarro, L., Alvarez-Manzaneda, I., de Vicente, I., 2016. 2016. Magnetic microparticles as a new tool for lake restoration: a microcosm experiment for evaluating the impact on phosphorus fluxes and sedimentary phosphorus pools. Water Res. 89, 366–374. Gonsiorczyk, T., Casper, P., Koschel, R., 2001. Mechanisms of phosphorus release from the bottom sediment of the oligotrophic Lake Stechlin: importance of the permanently oxic sediment surface. Arch. Hydrobiol. 151 (2), 203–219.
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Please cite this article as: Liu, X., et al., Effect of water quality improvement on the remediation of river sediment due to the addition of calcium nitrate, Sci Total Environ (2016), http://dx.doi.org/10.1016/j.scitotenv.2016.09.149