Environmental quality of Xuanwu Lake, China, after restoration

doi:10.21203/rs.3.rs-4102274/v1

Download PDF

Research Article

Environmental quality of Xuanwu Lake, China, after restoration

https://doi.org/10.21203/rs.3.rs-4102274/v1

This work is licensed under a CC BY 4.0 License

You are reading this latest preprint version

Some restoration measures were implemented in Xuanwu Lake to control water eutrophication, including sediment dredging, algal bloom control through clay flocculation, and the growth of aquatic macrophytes. The water quality of Xuanwu Lake was monitored before and after the restoration. The results showed that the sediment was not the primary pollution source of the water body, the dredging did not change the nutrient contents in the lake significantly, and the heavy metal content in the sediment also did not decrease significantly. Therefore, sediment dredging should be carefully selected; otherwise, aquatic ecosystems may be damaged, and exogenous pollution will become severe. It is necessary to restore aquatic plants after dredging. The water quality improved with increased water diversion and improved less when it exceeded a certain level. The volume of water diversion should be controlled at an appropriate level. Clay flocculation effectively inhibited cyanobacterial blooms within a short time, improved water quality, and promoted the restoration of the Potamogeton crispus population. Clay flocculation is an effective way to control water eutrophication under emergent conditions. In addition, water quality was improved distinctly during the growing period of the P. crispus population, compared to when aquatic vegetation was minimal. Aquatic population restoration is a sustainable and effective approach for managing Xuanwu Lake and controlling water eutrophication.

Algal bloom control

aquatic population

ecological restoration

sediment dredging

water diversion

Water eutrophication is one of the most critical important environmental problems worldwide (Luo et al. 2017; Zhang et al. 2022). Many measures have been taken to treat eutrophic water bodies, such as sediment dredging and water replacement, clay flocculation, and phytoremediation. Dredging is recognized as an effective way to improve water quality in eutrophic lakes and reduce contaminated sediment pollutants, especially endogenous nutrients and heavy metals (Reddy et al. 2007; James and Pollman 2011; Wen et al. 2020). The dredging can reduce internal nitrogen (N) and phosphorus (P) loads in some shallow lakes (Reddy et al. 2007; Yu et al. 2016) and metal concentrations in sediment. In an urbanized Florida bayou, the metal concentrations decreased by averaging 4–65% after dredging, especially for mercury (Hg), zinc (Zn), and chromium (Cr) (Lewis et al. 2001). However, dredging may bring a series of ecological problems, such as releasing contaminants from sediment into overlying water (Je et al. 2007) and decreasing benthic biodiversity and density (Lewis et al. 2001).

Clay flocculation is a valuable technique for controlling algal blooms in coastal and freshwater bodies (Hagström et al. 2010; Anderson et al. 2012). A modified clay treatment significantly reduced total phosphorus (TP) and total nitrogen (TN) in water (Lu et al. 2015). Clay flocculation effectively controlled algal blooms (Sengco and Anderson 2004). Modified clays could restrain 92% of the motility of algal cells in Taihu Lake, China, with an optimal clay concentration of 0.3 g/L, and it effectively controlled the Microcystis aeruginosa blooms in the lake (Liu et al. 2010). Similarly, the Cyanotoxin microcystins RR and LR in Taihu Lake were reduced by 50% and 40%, respectively, by a modified clay technology, and the submerged plants in the shallow lake were significantly restored with this technology (Pang et al. 2011).

The water diversion project effectively dilutes polluted water bodies and accelerates water exchange (Oseke et al. 2021). Theoretically, this technique could shorten renewal times and discharge nutrients from lakes to reduce water pollution (Tang et al. 2021). This project has already been applied to control water eutrophication in Taihu Lake, Chaohu Lake, and urban inland river treatment and has achieved some positive results (Hu et al. 2008; Tong et al. 2013; Tang et al. 2021). Nevertheless, this project is still controversial, as it may not solve the problem of nutrient over-enrichment and water quality degradation in lakes in the long run, and some adverse effects have even been reported in previous case studies (Khorasani et al. 2018; Qin et al. 2019; Yao et al. 2018). Water diversion positively impacts the ecological environment in some areas of Taihu Lake. However, it can only be used as an emergency measure to alleviate the hazards of blooms (Hu et al. 2008). If the concentration of phosphorus and nitrogen in lake water cannot be brought down to a manageable level, the risk of eutrophication in Taihu Lake will increase (Hu et al. 2008).

Aquatic plants are crucial markers of the health of aquatic ecosystems and water quality (Zhang et al. 2016). One efficient method of preventing water eutrophication and maintaining the stability of aquatic ecosystems is the restoration of aquatic vegetation, particularly submerged species (Qiu et al. 2001; Gao et al. 2017). Previous research (Qiu et al. 2001; Wang et al. 2017) demonstrated that the establishment of aquatic macrophytes enhanced water quality by reducing TN, TP, chemical oxygen demand, and total suspended particles while also increasing transparency. However, because of the low transparency and inadequate underwater light in eutrophic water, it is challenging to recover aquatic plants, especially submerged macrophytes (Wang et al. 2009).

Xuanwu Lake (32°4′N, 118°47′E) is located in the north of Nanjing City, Jiangsu province, and in the middle and lower reaches of the Yangtze River (MLYRA) of eastern China, where the economy is highly developed with a dense population (Fig. 1) (Cao et al., 2014). Xuanwu Lake is a small urban shallow lake (3.68 km²; mean depth 1.65 m; storage capacity 4.29 million m³). The lake water was eutrophic in the late1980s because of the rapid industrialization in the region. The sewage treatment and dredging projects were implemented to improve the lake's water quality from 1996 to 1998 (Zhu et al. 2004). After the dredging project in 1998, 8×10⁴ m³/d of water from the Yangtze River was diverted into Xuanwu Lake; in 2002, the water diversion was expanded to 18×10⁴ m³/d. Nevertheless, the cyanobacterial blooms, dominated by Microcystis, occurred in Xuanwu Lake for the first time in July 2005, and the clay flocculation technology was adopted to control the blooms. Potamogeton crispus appeared sporadically in the shallow parts of the southeast Lake and North Lake areas in November 2005, as cyanobacterial blooms were controlled and the P. crispus population started to grow and spread rapidly with increasing water temperatures. Since then, the P. crispus population has appeared in Xuanwu Lake annually in winter and flourished from spring to summer. Since 1998, the Yangtze River water has been introduced into Xuanwu Lake as its primary water source, then into the urban river channel of Nanjing, and the water diversion volume has increased from 8×10⁴m³/d to 28×10⁴m³/d. Therefore, the water quality of Xuanwu Lake is related to the health of the entire river ecosystem in the Nanjing urban area. Can the above measures improve the water quality of Xuanwu Lake? However, the impacts of these measures on the water quality of Xuanwu Lake still need to be clarified. This study attempted to compare the effectiveness of these measures in remediating water quality based on monitoring data. The information should have implications for the management of freshwater systems elsewhere.

Background

Sediment dredging

The Dredging project was done in January 1998 to remove about 30 cm of sediment from the lake, covering an area of 3.3 km² with a total volume of 87×10⁴ m³. The sediment formed in the past 40-100 years was removed according to the measured sedimentation rate of 0.3-0.7 mm/yr. On March 7, 1998, the project on the north lake was finished, and the southeast lake was also by the end of April. Pumping water from the Yangtze River replaced the lake water in late October 1998 and was finished by the end of 1998 (Fig. 2).

The grab dredger removed 40-50cm of sediment on December 27, 2008. The preliminary project was carried out in the north lake and completed in March 2009, with 68×10⁴ m³ removed. Subsequently, it was carried out in the southeast and southwest lakes. The dredging work was carried out in the southwest and southeast lakes, with a total water area of 81×10⁴ m² in 2014; the final dredging volume reached 43×10⁴ m³. The project was implemented in the southeast lake, with a dredging volume of 3×10⁴m³ in 2017, 9.3×10⁴m³ in 2018, 9×10⁴m³ in 2019, and 14.2×10⁴m³ of sediment in 2020.

Water diversion from the Yangtze River

Since the dredging project was completed in 1998, 8×10⁴ m³ of water has been diverted daily to Xuanwu Lake from the Yangtze River. By October 2002, when the second water diversion project was completed, the daily capacity of water diversion had increased to 18×10⁴m³. The cyanobacteria blooms broke out in Xuanwu Lake in July 2005, and the daily capacity increased to 28×10⁴m³ to control the blooms. Four water outlets introduced the lake water into urban rivers (Fig.1-b), and the annual water replenishment volume was about 50 million m³. The entire lake water can be replaced every 16 days. The changes in water diversion and the dredging project are listed in Fig. 2.

Emergency management with clay flocculation and P. crispus population recovery stage

The first recorded cyanobacteria blooms in Xuanwu Lake occurred in July 2005, with Microcystis dominating the blooms. On September 20, September 30, and October 16, 2005, algal blooms in the north, southeast, and southwest lakes were managed using a clay flocculation technique. The last day of treatment was October 31, 2005. About 300 tons of modified clay (kaolin) were applied for the treatment at a rate of 106 g/m².

The turions of P. crispus began to sprout naturally in the shallow waters before November 2005, after clay flocculation, and the seedlings of P. crispus were scattered sporadically in shallow water. The average population density was 3-5 plants/m², and the plant height ranged from 3 to 10 cm. Although P. crispus was still limited to shallow water from December 2005 to February 2006, plant height increased to 30-70 cm, with an average density of 30 plants per m². As the water temperatures rose rapidly in March, P. crispus flourished and expanded to the whole lake. In April 2006, the P. crispus plants were clipped artificially because of overpropagation. The top of the plants was cut off at about 20-30 cm, and the harvest was completed in mid-April. The life cycle of this plant species in the lake includes a sprouting stage at the end of October, a sporadic growth stage from November 2005 to February 2006, a rapid growth stage from March to April 10th, and the P. crispus population was harvested from the middle to the end of April.

Other measures

The amount of non-point source and point source pollution entering Xuanwu Lake decreased concurrently with the closure of the pollution sources surrounding the lake. Additionally, dredging has been done in a few of the lake's input rivers.

Sampling and analysis methods

We collected water samples jointly with the Nanjing Environmental Monitoring Center from 2003 to 2006, so our sampling sites were consistent, and the analysis results showed that our monitoring results were the same. The monitoring indicators include total nitrogen (TN), ammonium nitrogen (NH₃-N), total phosphorus (TP), permanganate index (PI), and five-day biochemical oxygen demand (BOD₅). Therefore, the annual water quality index from 1998 to 2020 was based on the data from the Nanjing Environmental Monitoring Center and the specific month data from our monitoring.

Five sampling points are shown in Fig. 1-b. Water temperature, dissolved oxygen (DO) concentration, and pH were measured using a multiparameter meter (YSI 6820, Yellow Spring Instruments, USA) at 9 o'clock, and Secchi disc transparency (SD) was measured using a standard Secchi disk. Water samples were collected at a depth of 5 cm below the surface. The water samples were pre-treated (5% potassium sulfate solution, 121 ºC, 30 min). TN, NH3-N, and TP were determined using an automated wet chemistry analyzer (SAN++, Skalar, Holland). Chlorophyll a content (Chla) was determined by the spectrophotometer method using acetone as extraction, and PI was determined by a titrimetric method by acid digestion with potassium permanganate oxidation (SEPA, 2002). The samplings were conducted at intervals of 7 days with three replicates.

The data on heavy metal concentration (1997–2000) was referenced from Zhu's research in 2004 (Zhu et al. 2004), and the data on heavy metals (2006–2020) was collected from the Nanjing Environmental Quality Report (by the Nanjing Environmental Monitoring Center). The data from the same monitoring points and the same monitoring methods are adopted to ensure the consistency of the data.

The Yangtze River water was precipitated by two waterworks and poured into Xuanwu Lake. The water quality parameters of the Yangtze River were used to compare the correlation with the water quality of Xuanwu Lake because the water quality parameters of the two waterworks were lacking. The water quality indicators of the Yangtze River (Nanjing section) were obtained from the Nanjing Statistical Yearbooks, respectively. The indicators include PI, BOD₅, NH₃-N, and TP, but TP is from 2003 to 2020; other indicators are from 1999 to 2020.

Geoaccumulation index (I_geo)

The geoaccumulation index (I_geo), designed by Müller (1969), has been widely used to study heavy metal pollution in sediments and soils. The geoaccumulation index (I_geo) is calculated by the following equation:

I_geo= log₂(Cn/1.5B_n) (1)

where Cn represents the heavy element's measured content (mg·kg^-1), and B_n represents the local geochemical background value of the heavy element (mg·kg^-1). The constant, 1.5, was the background matrix correction factor due to the lithogenic effect (Masocha et al. 2022; Mao et al. 2022).

The contamination level is classified into seven classes, where: Class 0: I_geo≤0 uncontaminated; Class 1: 0< I_geo≤1 uncontaminated to moderately contaminated; Class 2: 1 < I_geo≤2 moderately contaminated; Class 3: 2 < I_geo≤3 moderately to heavily contaminated; Class 4: 3 < I_geo≤4 heavily contaminated; Class 5: 4< I_geo≤5 heavily to extremely contaminated; and Class 6: I_geo >5 extremely contaminated (Masocha et al. 2022; Mao et al. 2022).

Statistical analyses

The data were subjected to variance analysis using SPSS 16.0. The significance of differences between treatments was determined at the 0.05 probability level (p<0.05).

Effects of sediment dredging and water diversion on lake

Water quality

The water quality parameters before and after dredging and water diversion are shown in Fig. 3-a. The annual average TN concentration in water was higher than 4.0 mg/L before dredging, decreased to the lowest level (2.47 mg/L) in 1999 after dredging, but increased to 4.87 mg/L in 2000. TN content in 2000 was higher than in 1997, before the dredging project and water diversion. The dredging projects were carried out in various lake areas of Xuanwu Lake from 2010 to 2020, and the rivers entering the lake were also intercepted or dredged simultaneously. The pollutants flowing into the lake have decreased, and the average TN decreased to 1.62 mg/L in 2010–2020, but it was still as high as 2.26 mg/L in 2011, exceeding the highest value in 2006–2009. The capacity of water diversion was 8×10⁴m³/d from 1999–2002, and the average yearly TN concentration was 2.96 mg/L. The capacity of water diversion was 18×10⁴ m³/d in 2003–2005, and the annual average TN concentration was 1.64 mg/L (1.48–1.83 mg/L). The P. crispus population was harvested in large numbers and removed from the lake in April 2006. Large amounts of nutrients such as nitrogen and phosphorus were carried from the water, and the capacity of water diversion was increased to 28×10⁴m³/d from 2006 to 2020. The TN content was slightly higher than when the water diversion amount was 18×10⁴m³/d but 72.1% lower than when the diversion volume was 8×10⁴m³/d (Table 1).

Table 1

Water quality parameters under different volume of water diversion
Parameters	Capacity of water diversion
	8×10⁴m³/d		18×10⁴m³/d		28×10⁴m³/d
	mean value	max	mean value	max	mean value	max
TN(mg·L^− 1)	2.96 ± 1.29	4.87	1.64 ± 0.18	1.84	1.72 ± 0.40	2.26
NH₃-N(mg·L^− 1)	0.87 ± 0.82	1.88	0.25 ± 0.05	0.28	0.20 ± 0.07	0.39
TP(mg·L^− 1)	0.30 ± 0.05	0.36	0.16 ± 0.02	0.18	0.09 ± 0.02	0.12
PI(mg·L^− 1)	7.04 ± 0.22	7.23	4.20 ± 0.24	4.46	3.27 ± 0.34	4.10
BOD(mg·L^− 1)	8.35 ± 1.60	10.70	3.89 ± 0.41	4.37	3.43 ± 0.86	4.90

The annual average NH₃-N concentration was 0.40 mg/L in 1997 before dredging and decreased to the lowest level (0.25 mg/L) in 1999 after dredging and water diversion. Although the water was replaced by the Yangtze River water for two years in 2000, the NH₃-N concentration was increased to 0.36 mg/L. The annual average NH₃-N concentration was 0.87 mg/L (0.13–1.88 mg/L) from 1999 to 2002, and the capacity of water diversion was 8×10⁴m³/d. When the capacity of water diversion increased to 18×10⁴m³/d from 2003 to 2005, the annual average NH₃-N concentration was 0.25 mg/L (0.19–0.28 mg/L). When the capacity of water diversion increased to 28×10⁴m³/d from 2006 to 2020, the average yearly NH₃-N concentration was 0.20 mg/L (0.13-0.39mg/L). When the capacity of water diversion increased from 18×10⁴m³/d to 28×10⁴m³/d, the average NH₃-N concentration decreased slightly, but the maximum value increased.

The annual mean TP concentration was 0.40 mg/L before dredging but decreased to 0.25 mg/L in 1999 and rose again to 0.36 mg/L in 2000. TP content remained higher level after dredging when the water diversion was 8×10⁴m³/d, and TP concentration in 2000 was 0.36 mg/L, and higher than the IV Classes of Water Standards in China (TP ≤ 0.3 mg/L) and near the V Classes of Water Standards (TP ≤ 0.4 mg/L). The average TP concentration was 0.30 mg/L (0.26–0.36 mg/L) in 1999–2002 when the capacity of water diversion was 8×10⁴m³/d. The average TP concentration decreased to 0.16 mg/L (0.14–0.18 mg/L) when the capacity of water diversion increased to 18×10⁴m³/d from 2003 to 2005. When the capacity of water diversion increased to 28×10⁴m³/d from 2006 to 2020, the average TP concentration decreased to 0.09 mg/L (0.07–0.12 mg/L).

PI decreased continuously from 1997 to 2000 after dredging, and PI in 2000 decreased by 45.3% compared to that in 1997 before dredging, but the average PI increased to 7.21 mg/L in 2002. The average PI was 7.04 mg/L (6.78–7.23 mg/L) in 1999–2002 when the capacity of water diversion was 8.0×10⁴m³/d, and the average PI decreased to 4.20 mg/L (3.99-4.46mg/L) when the capacity was 18×10⁴m³/d in 2003 to 2005. When the capacity of water diversion increased to 28×10⁴m³/d from 2006 to 2020, the average PI decreased to 3.27 mg/L (2.10-4.10mg/L), and PI showed an increasing trend from 2006 to 2020 and reached the highest value of 4.10 mg/L in 2020.

The concentration of BOD₅ was 15.91 mg/L in 1997 before dredging and decreased to the lowest level (7.96 mg/L) in 1999 after dredging and water diversion. However, the concentration of BOD₅ increased to 10.70 mg/L in 2000 and was higher than that in 1998. Although the Yangtze River has replaced the lake water for two years, it also exceeded the V Classes of Water Standards (BOD₅ ≤ 10.0 mg/L). Therefore, the water quality was still poor after dredging and water diversion of 8.0×10⁴m³/d. The average concentration of BOD₅ was 8.35 mg/L (7.19–10.70 mg/L) in 1999–2002 when the capacity of water diversion was 8.0×10⁴m³/d, and the average concentration decreased to 3.89 mg/L (3.64-4.37mg/L) when the capacity of water diversion increased to 18×10⁴m³/d in 2003–2005. When the capacity of water diversion was 28×10⁴m³/d from 2006 to 2020, the average concentration of BOD₅ decreased to 3.43mg/L (2.09-4.90mg/L), but the maximum value was higher than the peak in 2003–2005.

Sources of nutrients

Since the Yangtze River water was injected into Xuanwu Lake after precipitation, the correlation between the water quality indexes of Xuanwu Lake and the Yangtze River was analyzed (Table 2). It showed that NH₃-N concentration has a significant positive correlation between Xuanwu Lake and Yangtze River from 1999 to 2020 (r = 0.43, 95% level), and it also had a significant positive correlation from 2003 to 2020 (r = 0.48, 95% level). TP concentration did not have a significant correlation from 2003 to 2020 (r=-0.07, 95% level) but had a positive correlation from 2006 to 2020 (r = 0.48, not significant at the 95% level). The results showed that the nitrogen and phosphorus contents in Xuanwu Lake were significantly affected by the capacity of water diversion from the Yangtze River. There was no significant correlation between PI and BOD in Xuanwu Lake and the Yangtze River during different water diversion periods.

Table 2

The correlation between the water quality indexes of Xuanwu Lake and Yangtze River
Index Year	NH₃-N (95% level)			TP (95% level)		PI (95% level)			BOD₅ (95% level)
Index Year	1999–2020 (n = 22)	2003–2020 (n = 18)	2006–2020 (n = 15)	2003–2020 (n = 18)	2006–2020 (n = 15)	1999–2020 (n = 22)	2003–2020 (n = 18)	2006–2020 (n = 15)	1999–2020 (n = 22)	2003–2020 (n = 18)	2006–2020 (n = 15)
correlation coefficient	0.43*	0.48*	0.48	-0.07	0.37	-0.12	-0.34	-0.32	-0.12	-0.31	0.29

The N and P contents of influent and effluent in 2006 and 2010 are shown in Table 3. The results showed that the N carried by water diversion accounted for 65.7% of TN into Xuanwu Lake, and the P was 46.5%. The release of N and P from sediment only accounted for 21.6% and 19.7%, respectively. When the capacity of water diversion increased to 28×10⁴m³/d, the release of N and P from sediment only accounted for 31.0% and 20.9%, respectively, in 2010, mainly contributed by water diversion.

Table 3

Nitrogen and phosphorus content inlet and outlet of Xuanwu Lake in 2006 and 2010
Year(capacity of water diversion)	Component	Input/tons						output/tons
Year(capacity of water diversion)	Component	Precipitation	Runoff	Point source	Ditches and pipes	Water diversion	Sediment Release	River	P. crispus harvest
2006 (18×10⁴m³)	TN	9.77^a	3.25^a	1.64^a	53.43^a	130.15^a	42.83 ^b	90.38^a	65.6^a
2006 (18×10⁴m³)	TP	0.13^a	0.28^a	0.18^a	2.87^a	4.75^a	2.01^c	4.76^a	1.5^a
2010 (28×10⁴m³)	TN	20.72^d	/	/	/	163.24^e	50.54 ^b		/
2010 (28×10⁴m³)	TP	0.69^d	/	/	/	7.42^e	1.55 ^b		/
a Obtained from Nanjing Environmental Quality Report (2006–2010); b, c Calculated according to Gong etal’s (2006; 2007) research; d Calculated according to Sun etal’s (2014) research; e Calculated according to the actual data measured by Nanjing Municipal Water Bureau

Therefore, N and P from sediment were not the primary sources in lakes, and the dredging did not improve the water quality. When the capacity of water diversion increased from 8×10⁴m³/d to 18×10⁴m³/d, the water quality was improved. However, when the water diversion increased to 28×10⁴m³/d, the improvement of water quality was the contribution of the P. crispus harvest in 2005 and the interception of pollution sources around the lake, not the contribution of the increase in water diversion.

Effects of dredging and water diversion on heavy metals in sediments

The heavy metal contents in sediments before and after dredging and water diversion are shown in Fig. 4. The heavy metals content was higher before dredging in 1996 and 1997; the content of As, Hg, Cr, Pb, and Cd decreased to the lowest level in the dredging period (1998); but the content of Cu and Zn in 1998 was similar to that in 1997, and all heavy metals content continued to increase after dredging in 1999 and 2000. The content of Zn in 1999 reached the highest value, and the content of Hg, Pb, and Cu in 2000 reached the highest value. The content of other heavy metals in 2000 was higher than in 1998.

It was found that the amount of water diversion had no significant effect on the heavy metal content in sediments. The content of all heavy metals reached a relatively high level during 2010–2013; however, the amount of water diversion in this period was higher than that from 1999 to 2005. In particular, the content of As was 23.4 mg/kg in 2013, 349% higher than that in 1999; the content of Hg, Cd, and Cu in 2013 also reached a high level.

Although the volume of water diversion increased by 10⁴m³/d from 2003 to 2005, the average contents of As, Cr, Cd, and Cu were higher than those in 1999–2000; only Hg, Pb, and Zn decreased. When the water diversion volume increased by 28×10⁴m³/d from 2006 to 2020, the average contents of As, Cr, and Cd were still higher than those in 1999–2000, and the content of Cu had no significant difference that in 1999–2000, and the contents of Hg, Pb, and Zn had little difference from those in 2004–2005.

The I_geo of heavy metals in sediment is listed in Fig. 5. I_geo of As was always less than 0 or between 0 and 1 from 1996 to 2020; the contamination level was classified into Class 0 or Class 1, but I_geo of As reached the highest value (0.55) in 2013. The I_geo of Hg was mainly between 1 and 2 from 1996 to 2020; the contamination level was classified into Class 2. The I_geo of Hg in 2000 was 2.81, reaching Class 3. The I_geo of Cr, Pb, and Cu was mainly less than 0 or between 0 and 1 from 1996 to 2020; the contamination level was classified into Class 0 or 1. The I_geo of Cr reached the highest value of 0.54 in 2010; the I_geo of Pb reached the highest value of 0.68 in 2000, and the I_geo of Cu reached the highest value of 1.09 in 2004. The I_geo of Cd was mainly less than 0 or between 0 and 1 from 1996 to 2020; the contamination level was classified into Class 0 or Class 1, but more than 1 in 2004, 2005, and 2013, and reached the highest value (1.17) in 2005. The I_geo of Zn was between 2 and 3 from 1997 to 1999 and decreased to 0–2 from 2000 to 2020; the contamination level decreased slightly. The highest value of I_geo of As and Cr occurred in the period of maximum water diversion, and the I_geo of Hg, Pb, Cd, and Cu all was at a high level from 2010 to 2013; only the I_geo of Zn decreased after 2000 compared with that before 2000. These indicate that the amount of water diversion and dredging had little effect on the content of heavy metals in sediments and the impact of emergency treatment with clay flocculation and P. crispus population growth and harvesting on water quality.

Although the nutrient level was low in 2005 compared with other years, the cyanobacteria blooms still broke out in July 2005, so clay flocculation was used to control the algal blooms. The water quality indexes before and after clay flocculation treatment are shown in Table 4. SD was only 20 cm before treatment with clay flocculation, increased to 41 cm significantly during the treatment period by 103% (p < 0.05), and further to 56 cm after the treatment, an increase of 1.8 times (p < 0.05). DO concentration was 7.67 mg/L before clay flocculation, increased to 8.73 mg/L during the treatment, and further to 9.52 mg/L after the treatment, but the difference was not statistically significant before and after treatment (p > 0.05). TN concentration was 4.59 mg/L before the treatment, significantly decreased to 1.59 mg/L during the treatment (p < 0.05), and slightly increased to 1.66 mg/L after clay flocculation, a decrease of 63.9% before and after the treatment. TP concentration was 1.90 mg/L before the treatment, decreased to 0.70 mg/L during the treatment, and decreased to 0.25 mg/L after the treatment, decreased by 86.8%, which was significant before and after treatment (p < 0.05).

Table 4

Effect of clay flocculation and population restoration on water quality
Measures	Phase	DO(mg·L^− 1)	SD(cm)	TN(mg·L^− 1)	TP(mg·L^− 1)
Clay flocculation	Before treatment	7.7	20.0	4.6	1.90
	During treatment	8.7	41.0	1.6	0.70
	After treatment	8.7	56.0	1.7	0.25
	Changing rate(%)	13.8	180.0	63.9	86.8
Population restoration	Before restoration	8.8	44.5	3.0	0.25
	Sporadic growth stage	10.2	92.6	2.2	0.14
	Rapid growth stage	10.1	99.4	2.4	0.12
	Harvest stage	6.8	92.0	1.2	0.06
	Changing rate(%)	23.7	106.7	58.3	77.9

P. crispus proliferated from November 2005 to March 2006 because of the improvement in water quality, and spread to the whole lake in April 2006, and then was harvested. The water quality indexes before and after the restoration of the P. crispus population are shown in Table 4. DO in water was the highest at the sporadic growth stage and the rapid colonization stage of P. crispus, reaching 10.2 mg/L and 10.09 mg/L, respectively, higher than that before restoration, but decreased to 6.75 mg/L at the harvest stage, and increased 23.7% than that before restoration. SD in Xuanwu Lake increased gradually with the restoration of the P. crispus population, from sporadic growth to the colonization stage, but decreased after the plants were harvested. For instance, SD was only 44.5 cm in the pre-recovery stage, increased to 92.6 cm at the sporadic growth stage, to 99.4 cm at the colonization stage, but decreased to 92.0 cm after the harvest stage. TN concentration in water was 2.97mg/L at the pre-recovery stage, decreased to 2.23 mg/L and 2.36 mg/L, respectively, at the sporadic growth and colonization stage, and decreased rapidly to 1.24 mg/L at the harvest stage, which was 58.3% lower, as compared to the pre-recovery stage. TP in water was the highest (0.25 mg/L) at the pre-recovery stage, decreased to 0.14 mg/L and 0.12 mg/L, respectively, at the sporadic growth and colonization stage, and further to 0.06 mg/L at the harvest stage, which was more than four times lower as compared to the pre-recovery stage.

Because P. crispus was harvested in a large area, it carried a large amount of N and P from the water, decreasing rapidly. The total amount of N and P entering the lake in Xuanwu Lake was 198.24 t/a and 8.21 t/a, respectively, and the total amount of N and P out of the lake was 155.98 t/a and 6.26 t/a, respectively (Table 3). 78.7% of the N and 76.2% of the P were carried out of the Xuanwu Lake through harvesting. Therefore, TN and TP in the water decreased significantly during the harvest stage.

The water quality indexes in the same months (September-October) before and after a few years are listed in Table 5. After the clay flocculation, SD was not only improved in 2005 but also decreased compared with that in the same period of other years. There was no significant difference before and after clay flocculation. SD was 45cm, 50cm, and 58cm (mean = 47.5cm) from September to October 2004; SD was only 42cm, 50cm, and 56cm (mean = 46.0cm) in 2006; and there was no significant difference in SD between 2004 and 2006.

Table 5

Comparison of clay flocculation and population restoration on water quality in the same period
Measures	Year	SD(cm)	TN(mg·L^− 1)	TP(mg·L^− 1)	Chla(µg·L^− 1)
Clay flocculation	2003(Sep.-Oct.)	26.5	1.78	0.15	60
	2004(Sep.-Oct.)	47.5	1.55	0.19	93
	2005(Sep.-Oct.)*	32.0	2.06	0.27	214
	2006(Sep.-Oct.)	46.0	1.60	0.19	50
Population restoration	2003(Mar.-Jun.)	33.8	1.28	0.13	50
	2004(Mar.- Jun.)	28.3	1.39	0.16	41
	2005(Mar.- Jun.)	35.5	1.85	0.15	11
	2006(Mar.- Jun.)*	81.0	1.20	0.08	29
* Time of implementation of clay flocculation and population restoration

TN and TP in 2005 (Sep. to Oct.) were the highest in the same period from 2003 to 2006. TN showed no significant improvement after the clay flocculation; TN from September and October 2006 was similar to that in 2003 and 2004. TP in 2006 (Sep. to Oct.) was identical to 2004, but there was no significant improvement compared to 2003. Chla content in 2005 was much higher than that in other years. So, clay flocculation did not improve the water quality of Xuanwu Lake, and the algal bloom has yet to be controlled entirely. The water quality indexes between the same period in the growing season of the P. crispus population are listed in Table 5. The population of P.crispus expanded rapidly from March to April. Water transparency from March to June in 2006 was significantly higher than that from 2003 to 2005, especially over 186% higher than that from January to June of 2004. TN and TP content in March–June 2006 were all the lowest from 2003 to 2005. TP content in March–June 2006 was lower than in 2003–2005. TP content was lower than the II Classes of Water Standards (TP ≤ 0.1 mg/L). The Chla content in March–June 2006 was lower than that in 2003 and 2004 but higher than that in 2005; it indicated that P.crispus significantly improved water quality in the vigorous growth period compared with the same period in other years.

Effects of sediment dredging

It is crucial to control both external and endogenous pollution sources to prevent water eutrophication (Wang et al. 2015). Dredging is a commonly adopted means to control the release of endogenous pollutants (Zhang et al. 2014). However, the dredging also changed the underwater terrain, increased suspended matter, and destroyed the original population structure of organisms and habitat of aquatic ecosystems, especially the decomposer population and aquatic plants, thus weakening the self-purification ability of the lake (Wei 2012). The release of nutrients from the sediment to overlying water is related to temperature, pH, redox potential, bacteria, dissolved oxygen, and many other factors (Peng et al. 2021). The sediment at the bottom is not directly involved in releasing nutrients to overlying water; once the upper sediment is dredged, the bottom sediment will be exposed as surface sediment. At the overlying water-sediment interface of a healthy aquatic ecosystem, the residues of dead algae are decomposed by bacteria, and N is released back into the water body through a series of processes, such as ammonification and nitrification (Zhao et al. 2017). Therefore, the total nitrogen in the sediment can maintain balance (Zhao et al. 2017). Sediment dredging may destroy the original ecosystem of the overlying water-sediment interface, and the decomposition of dead algae is affected, which may be one of the reasons for the increase of TN after dredging. For the same reason, the dredging breaks the balance of the P cycle between the sediment and overlying water, resulting in an increase in TP concentration in the water (Wei 2012). It can also be seen from the analysis of Fig. 3 and Table 1 that the water quality indexes did not significantly improve after a series of dredging. Meanwhile, as can be seen from Table 3, the sediment release was not the primary source of N and P in the water. Therefore, the dredging had little influence on TN and TP contents.

The balance between bottom and surface sediment was disrupted when the surface sediment was removed. The heavy metals in the bottom sediment were exposed to achieve a new equilibrium, and the release flux of heavy metals in the bottom sediment was upward. It can be seen from Fig. 4 and Fig. 5 that the contents of all heavy metals in surface sediments were the same as before dredging after two years of release. The Hg, Pb, and Cu contents significantly exceeded those before dredging. Xuanwu Lake was polluted by heavy metals from the 1960s to the 1970s, and the heavy metal pollution in the sediments was relatively heavy. However, heavy metal pollution has been effectively controlled since the 1980s (Gong et al. 2006). Gong et al. (2006) found that the sediments of Xuanwu Lake are mainly clay particles after dredging, and the particle size is mostly less than 100 µm, more than 90%. Because the clay particles have a strong adsorption capacity for heavy metals (Tabatabaei et al. 2012). Moreover, with the increase in water diversion, the residence time of water in the lake became shorter, and the exchange time between sediment and overlying water became shorter. Therefore, the release of heavy metals in bottom sediment and the clay adsorption would reach a new equilibrium after dredging; the heavy metals in sediments remained relatively stable for a long time, with little change. So, the dredging project since 1998 and a series of subsequent projects had little effect on the heavy metals in the sediment.

Therefore, the restoration of hydrophytes should be carried out after dredging (Wei 2012). Nevertheless, the propagules of aquatic plants in Xuanwu Lake were also removed during the dredging treatment, so the lake lacked aquatic plants after the dredging. The lake ecosystem's inorganic environment, producers, and decomposers have changed. Therefore, water quality did not improve during the treatment and the extension of time after that.

Effects of water diversion

After sedimentation and filtration, the Yangtze River water was introduced into Xuanwu Lake, but the nutrient content was high (Zhang and Xu 2012). Because the water of the Yangtze River was flowing, the water velocity of Xuanwu Lake was slower than that of the Yangtze River, and the nutrient content would accumulate in the lake, resulting in a nutrient index in the lake was higher than that of the Yangtze River.

The storage capacity of Xuanwu Lake is low (4.3×10⁶ m³); when the capacity of water diversion was 18×10⁴m³/d, the lake water could be renewed once in about 25 days; if the capacity increased to 28×10⁴m³/d, the entire lake could be replaced about once every 16 days. According to the analysis of the nutrient content of water in different diversion periods, when the diversion volume increased from 8×10⁴m³/d to 18×10⁴m³/d, the water quality improved. One reason was that the water from the Yangtze River replaced the water in Xuanwu Lake with higher nutrients. Another important reason was that the non-point source pollution around Xuanwu Lake had been intercepted by the end of 2003. So, the nutrients flowing into the lake had been significantly reduced. However, TN improved when the water diversion increased to 28×10⁴m³/d, although TP content and PI in the water decreased. So, the water quality indexes did not change significantly and exceeded previous years. The critical reason for the decrease in TP and other indexes was the large-scale harvest of Potamogeton crispus in 2006, which carried many nutrients from the water.

The hydraulic retention time of lake water was shortened with the increase in water diversion from 2006 to 2020, and the exchange time of N and P between the overlaying water and sediment became shorter. From Table 3, P in the water mainly comes from water diversion, so the TP content in Xuanwu Lake was consistent with that in the Yangtze River. But N sources in the lake were more diverse: water diversion, precipitation, ditches and runoff point sources, and so on. Therefore, the N content in Xuanwu Lake exceeded that in the Yangtze River, but the N content was also highly correlated with the water in the Yangtze River. Zhang et al. (2012) measured the water quality indicators of Daqiao waterworks from 2005 to 2006 and compared them with the water quality indicators of Xuanwu Lake; they found a significant positive correlation between N content in waterworks and Xuanwu Lake; the P content showed a positive correlation but was not significant. Some studies showed that the water diversion of Xuanwu Lake only needs 15×10⁴m³/d -18×10⁴m³/d (Pang 2021), which can only meet the water quality needs. So when the amount of water diversion was 28×10⁴m³/d, which had little impact on water quality, and was too large, the lake gradually lost its ecological function, and the economic burden increased.

Effects of clay flocculation and aquatic macrophyte restoration

It can be shown in Table 4. After emergency treatment with modified clay, cyanobacteria were retained in the clay and precipitated on the sediment surface. In addition, the clay could adsorb suspended particles in water, leading to a notable improvement in SD and an increase in DO in the water column. The decrease in the concentrations of N and P in the water column may be attributed to the retention of ammonium and phosphate in the clay and the reduction of their release from sediment due to the settlement of fresh clay on the surface of sediments. According to Zhang (2006), zooplankton in 1 liter of water increased by 47%, 58% in protozoa, 14% in rotifers, 446% in Cladocera, and 180% in copepods. Fish activities were normal, including black carp (Mylopharyngodon piceus), grass carp (Ctenopharyngodon idellus), silver carp (Hypophthalmichthys molitrix), bighead carp (Aristichthys nobilis), and squid without the occurrence of dead fish. All these indicate that emergency treatment can effectively improve water quality (Zhang et al. 2006). Emergency management also provides favorable conditions for the emergence of a large number of P. crispus populations in the later period (Wang et al. 2017).

It can be seen in Table 4 and Table 5. During the growth of P. crispus, it can absorb nutrients, reduce water flow, and improve water SD. P. crispus releases oxygen through photosynthesis during its growth, which increases DO concentration in water (Chen et al. 2007). When P. crispus was harvested, the water body lost the function of aquatic plants, and DO concentration decreased. P. crispus consumed a large amount of N and P from the water (Xie et al. 2015; Wang et al. 2017), so the TN and TP concentrations in the water were significantly reduced at the growth stage than during the population recovery period. When P. crispus was harvested, a large quantity of N and P were removed from the water, and the amounts of TN and TP in the water body decreased.

Water flow velocity was reduced because of the damping effect of the P. crispus population, and the suspended solids in water were precipitated, which improved the transparency. For instance, water transparency in January-June 2006 was significantly higher than in previous years. P. crispus can absorb a large amount of N and P from the waterbody during its rapid growth stage (Xie et al. 2015; Wang et al. 2017). Consequently, N and P concentrations in water were also lower from March to June 2006 than in the previous year. As P. crispus was harvested, large amounts of N and P nutrients were removed from the water system, so the TN and TP concentrations from May to June were also lower than those in the previous year.

Dredging did not lead to the emergence of a massive P. crispus population since it destroyed aquatic habitat and removed aquatic plant propagates from the sediment. Therefore, water quality was improved shortly after dredging but declined in the long term, likely due to the absence of aquatic plant communities. Great efforts have been made to grow aquatic plants in Xuanwu Lake, but only P. crispus appears sporadically in shallow waters every year. Nevertheless, water transparency was improved after the emergency treatment by clay flocculation, and the underwater illumination was enhanced. The intensified underwater illumination promoted the turion germination of P. crispus, and the increase in dissolved oxygen and the decrease in nutrient concentration facilitated the growth of seedlings of P. crispus and the expansion of the population later.

Dredging reduced the concentration of pollutants and heavy metals in sediments and alleviated water eutrophication quickly, but there was no improvement in the long term. The water quality was improved by increasing the volume of water diversion from 8×10⁴m³/d to 18×10⁴m³/d, but increased to 28×10⁴m³/d; the improvement of water quality has been improved but not significantly, and the main reason for the improvement was that the harvest of P. crispus carried a large amount of nutrients from the lake in 2006. Therefore, the amount of water diversion should be controlled within an appropriate range (such as 18×10⁴m³/d), which can reduce the waste of water resources and investment. It is also necessary to establish buffer strips and ecological ditches around the lake to purify the water bodies entering the lake and increase the water supply.

The cyanobacteria blooms were effectively restrained by clay flocculation. The transparency and water quality were improved after the treatment. The favorable conditions were provided for the sprouting and growth of the P. crispus population. It is an effective way to enhance water quality in the short term. The effect of population restoration on improving lake water quality is more evident and persistent than that of the dredging project and clay flocculation. Therefore, in the process of treating lake eutrophication, ecological restoration should be carried out on the premise of controlling exogenous pollution. However, in the ecological restoration of Xuanwu Lake, the emergence of the P. crispus population would expand excessively and cause secondary pollution when the population mass declined. Therefore, it is necessary to harvest gradually at this stage, making the population expand continuously and taking out nutrients from the water.

Funding information

This work was supported by the National Natural Science Foundation of China (No. 42077303) under grant.

Contributions Jinqi Wang: Conceptualization, writing original draft, figures formation, tables formation; Yuzhi Song: methodology, revision and proofreading. All authors have read and agreed to the published version of the manuscript.

Availability of data and materials Not applicable.

Conflict of interest: The authors declare no competing interests.

Anderson DM, Cembella AD, Hallegraeff GM (2012) Progress in understanding harmful algal blooms: paradigm shifts and new technologies for research, monitoring, and management. Annual Rev Mar Sci 4:143–176. https://doi.org/10.1146/annurev-marine-120308-081121
Cao YM, Zhang E, Cheng G (2014) A primary study on relationships between subfossil chironomids and the distribution of aquatic macrophytes in three lowland floodplain lakes, China. Aquat Ecol 48:481–492. https://doi.org/10.1007/s10452-014-9499-7
Chen C, Yin D, Yu B, Zhu HK (2007) Effect of epiphytic algae on photosynthetic function of Potamogeton crispus. J Freshw Ecol 22:411–420. https://doi.org/10.1080/02705060.2007.9664171
Hagström JA, Sengco MR, Villareal TA (2010) Potential methods for managingPrymnesium parvum blooms and toxicity, with emphasis on clay and barley straw: a review. J Am Water Resour Assoc 46:187–198. https://doi.org/10.1111/j.1752-1688.2009.00402.x
Hu W, Zhai S, Zhu Z, Han H (2008) Impacts of the Yangtze River water transfer on the restoration of Lake Taihu. Ecol Eng 34:30–49. https://doi.org/10.1016/j.ecoleng.2008.05.018
Gao HL, Qian XH, Wu HF, Li HM, Pan H, Han CM (2017) Combined effects of submerged macrophytes and aquatic animals on the restoration of a eutrophic water body—A case study of Gonghu Bay, Lake Taihu. Ecol Eng 102:15–23. https://doi.org/10.1016/j.ecoleng.2017.01.013
Gong CS, Yao Q, Fang CX, Dong HP, Zheng CH, Bao XM (2006) Release fluxes estimate of phosphorus in a urban shallow lake:Lake Xuanwu, Nanjing. J Lake Sci 18:179–183 (in Chinese). https://doi.org/10.18307/2006.0212
Gong CS (2007) Study on small urban shallow lake inner pollution source and environmental dredging depth. Dissertation (in Chinese)
James RT, Pollman CD (2011) Sediment and nutrient management solutions to improve the water quality of Lake Okeechobee. Lake Reserv Manag 27:28–40. https://doi.org/10.1080/07438141.2010.536618
Je C, Hayes DF, Kim K (2007) Simulation of resuspended sediments resulting from dredging operations by a numerical flocculent transport model. Chemosphere 70:187–195. https://doi.org/10.1016/j.chemosphere.2007.06.033
Khorasani H, Kerachian R, Malakpour-Estalaki S (2018) Developing a comprehensive framework for eutrophication management in off-stream artificial lakes. J Hydrol 562:103–124. https://doi.org/10.1016/j.jhydrol.2018.04.052
Lewis MA, Weber DE, Stanley RS, Moore JC (2001) Dredging impact on an urbanized Florida bayou: effects on benthos and algal-periphyton. Environ Pollut 115:161–171. https://doi.org/10.1016/S0269-7491(01)00118-X
Liu GF, Fan CX, Zhong JC, Zhang L, Ding SM, Yan SH, Han SQ (2010) Using hexadecyl trimethyl ammonium bromide (CTAB) modified clays to clean the Microcystis aeruginosa blooms in Lake Taihu, China. Harmful Algae 9:413–418. https://doi.org/10.1016/j.hal.2010.02.004
Lu GY, Song XX, Yu ZM, Cao XH, Yuan YQ (2015) Environmental effects of modified clay flocculation on Alexandrium tamarense and paralytic shellfish poisoning toxins (PSTs). Chemosphere 127:188–194. https://doi.org/10.1016/j.chemosphere.2015.01.039
Luo L, Duan N, Wang XC, Guo WS, Ngo HH (2017) New thermodynamic entropy calculation based approach towards quantifying the impact of eutrophication on water environment. Sci Total Environ 603–604:86–93. https://doi.org/10.1016/j.scitotenv.2017.06.069
Masocha BL, Dikinya O, Moseki B (2022) Bioavailability and contamination levels of Zn, Pb, and Cd in sandy-loam soils. Botsw Environ Earth Sci 81:171. https://doi.org/10.1007/s12665-021-10129-3
Mao LC, Kong H, Li FP, Chen ZJ, Wang L, Lin T, Lu ZB (2022) Improved geochemical baseline establishment based on diffuse sources contribution of potential toxic elements in agricultural alluvial soils. Geoderma 410:115669. https://doi.org/10.1016/j.geoderma.2021.115669
Müller G (1969) Index of geoaccumulation in sediments of Rhine River. GeoJournal 2:108–118
Oseke FI, Anornu GK, Adjei KA, Eduvie MO (2021) Assessment of water quality using GIS techniques and water quality index in reservoirs affected by water diversion. Water-Energy Nexus 4:25–34. https://doi.org/10.1016/j.wen.2020.12.002
Pan G, Yang B, Wang D, Chen N, Tian BH, Zhang ML, Yuan XZ, Chen YJ (2011) In-lake algal bloom removal and submerged vegetation restoration using modified local soils. Ecol Eng 37:302–308. https://doi.org/10.1016/j.ecoleng.2010.11.019
Pang M, Song WW, Qian C (2021) Water quantity optimization method for improving water quality of Xuanwu Lake by water diversion. Water Resour Prot 37:133–139 (in Chinese)
Peng C, Huang YY, Yan XC, Jiang L, Wu XF, Zhang W, Wang XY (2021) Effect of overlying water pH, temperature, and hydraulic disturbance on heavy metal and nutrient release from drinking water reservoir sediments. Water Environ Res 93:2135–2148. https://doi.org/10.1002/wer.1587
Qin B, Paerl HW, Brookes JD, Liu J, Jeppesen E, Zhu G, Zhang Y, Xu H, Shi K (2019) Why Lake Taihu continues to be plagued with cyanobacterial blooms through 10 years (2007–2017) efforts. Chin Sci Bull 64:354–356. https://doi.org/10.1016/j.scib.2019.02.008
Qiu DR, Wu ZB, Liu BY, Deng JQ, Fu GP, He F (2001) The restoration of aquatic macrophytes for improving water quality in a hypertrophic shallow lake in Hubei Province, China. Ecol Eng 18:147–156. https://doi.org/10.1016/S0925-8574(01)00074-X
Reddy KR, Fisher MM, Wang Y, White JR, Thomas JR (2007) Potential Effects of Sediment Dredging on Internal Phosphorus Loading in a Shallow, Subtropical Lake. Lake Reserv Manag 23:27–31. https://doi.org/10.1080/07438140709353907
Sengco MR, Anderson DM (2004) Controlling harmful algal blooms through clay flocculation. Eukaryot Microbiol 51:169–172. https://doi.org/10.1111/j.1550-7408.2004.tb00541.x
State Environmental Protection Administration (SEPA) (2002) Monitoring and Analysis Methods for Water and Wastewater, 4th edn. China. Environmental Science Press(in Chinese, Beijing
Sun LY (2014) Atmospheric nitrogen and phosphorus deposition nn nanjing and effects of simulated nitrogen deposition on gaseous emissions of soils. Dissertation (in Chinese)
Tabatabaei SH, Najafi P, Mirzaei S, Nazem Z, Heidarpour M, Hajrasoliha S, Afyuni M, Harchegani HB, Landi E, Akasheh L (2012) Compost' leachate recycling through land treatment and application of natural Zeolite. Int J Recycling Org Waste Agric 1:2. https://doi.org/10.1186/2251-7715-1-2
Tang CY, He C, Li YP, Acharya K (2021) Diverse responses of hydrodynamics, nutrients and algal biomass to water diversion in a eutrophic shallow lake. J Hydrol 593:125933. https://doi.org/10.1016/j.jhydrol.2020.125933
Tong CF, Lv LR, Shao YY, Hao JL (2013) Calculation and analysis of water quality improvement caused by water diversion in Nanjing Qinhuai River. Adv Civil Eng II 256–259:2528–2532. https://doi.org/10.4028/www.scientific.net/AMM.256-259.2528
Wang GX, Zhang LM, Chua H, Li XD, Xia MF (2009) A mosaic community of macrophytes for the ecological remediation of eutrophic shallow lakes. Ecol Eng 35:582–590. https://doi.org/10.1016/j.ecoleng.2008.06.006
Wang JQ, Song YZ, Wang GX (2017) Causes of large Potamogeton crispus L. population increase in Xuanwu Lake. Environ Sci Pollut Res 24:5144–5151. https://doi.org/10.1007/s11356-016-6514-7
Wang TC, Qu GZ, Sun QH, Liang DL, Hu SB (2015) Evaluation of the potential of p-nitrophenol degradation in dredged sediment by pulsed discharge plasma. Water Res 84:18–24. https://doi.org/10.1016/j.watres.2015.07.022
Wei HU (2012) Study on key problems of a new environmental dredging based on ecological protection and subsequent ecological restoration. Meteorological Environ Res 11:47–49
Wen S, Zhong J, Li X, Zhang YL, Wang CH, Zhang L (2020) Does external phosphorus loading diminish the effect of sediment dredging on internal phosphorus loading? An in-situ simulation study. J Hazard Mater 394:122548. https://doi.org/10.1016/j.jhazmat.2020.122548
Xie D, Zhou HJ, Zhu H, Ji HT, Li N, An SQ (2015) Differences in the regeneration traits of Potamogeton crispus turions from macrophyte and phytoplankton-dominated lakes. Sci Rep 5:12907. https://doi.org/10.1038/srep12907
Yao X, Zhang L, Zhang Y, Du Y, Jiang X, Li M (2018) Water diversion projects negatively impact lake metabolism: a case study in Lake Dazong, China. Sci Total Environ 613–614:1460–1468. https://doi.org/10.1016/j.scitotenv.2017.06.130
Yu JH, Fan CX, Zhong JC, Zhang YL, Wang CH, Zhang L (2016) Evaluation of in situ simulated dredging to reduce internal nitrogen flux across the sediment-water interface in Lake Taihu, China. Environ Pollut 214:866–877. https://doi.org/10.1016/j.envpol.2016.03.062
Yu M, Wang C, Liu Y, Olsson G, Wang C (2018) Sustainability of mega water diversion projects: Experience and Lessons from China. Sci Total Environ 619:721–731. https://doi.org/10.1016/j.scitotenv.2017.11.006
Zhang R, Zeng FX, Liu WJ, Zeng RJ, Jiang H (2014) Precise and economical dredging model of sediments and its field application: case study of a river heavily polluted by organic matter, nitrogen, and phosphorus. Environ Manage 53:1119–1131. https://doi.org/10.1007/s00267-014-0268-0
Zhao SN, Shi XH, Li CY, Zhang S, Sun B, Wu Y, Zhao SX (2017) Diffusion flux of phosphorus nutrients at the sediment–water interface of the Ulansuhai Lake in northern China. Water Sci Technol 75:1455–1465. https://doi.org/10.2166/wst.2017.017
Zhang YL, Liu XH, Qin BQ, Shi K, Deng JM, Zhou YQ (2016) Aquatic vegetation in response to increased eutrophication and degraded light climate in Eastern Lake Taihu: Implications for lake ecological restoration. Sci Rep 6:23867. https://doi.org/10.1038/srep23867
Zhu M, Wang G, Wang J, Chen C (2004) Comparative analysis of changes of pollutants in sediment in Nanjing Xuanwu lake before and after sediment dredging. J Nanjing Normal Univ (Engineering Technol Edition) 42:66–69
Zhang X, Zhao J, Ding L, Li Y, Liu HX, Zhao YF, Fu G (2022) Eutrophication evolution trajectory influenced by human activities and climate in the shallow Lake Gehu, China. Ecol Ind 138:108821. https://doi.org/10.1016/j.ecolind.2022.108821
Zhang ZH, Xu Y (2012) Effects analysis on water quality of ecological water compensation from yangtze river to Xuanwu Lake. E Environ Monit Manage Technol 20:40–43

Download PDF

Reviewers agreed at journal
01 Jul, 2024
Reviewers invited by journal
03 Apr, 2024
Editor invited by journal
25 Mar, 2024
Editor assigned by journal
24 Mar, 2024
First submitted to journal
24 Mar, 2024
Editorial decision: Major revisions
16 Mar, 2024

You are reading this latest preprint version

Environmental quality of Xuanwu Lake, China, after restoration

Status:

Version 1

Abstract

Figures

Introduction

Materials and Methods

Background

Sampling and analysis methods

Statistical analyses

Results

Effects of sediment dredging and water diversion on lake

Water quality

Sources of nutrients

Effects of dredging and water diversion on heavy metals in sediments

Discussion

Effects of sediment dredging

Effects of water diversion

Effects of clay flocculation and aquatic macrophyte restoration

Conclusions

Declarations

References

Status:

Version 1