Introduction

Freshwater sediments are depositions that accumulate at the bottom of water bodies such as rivers, lakes, and reservoirs for a long time. They are an important part of a multi-phase ecosystem of water bodies. However, sediment deposition in a watershed reduces flood storage capacity. Soil erosion in catchments has long been an important cause of freshwater deposition. In turn, soil erosion is caused by anthropogenic disturbances such as deforestation, excessive agriculture, and mining. The sediment delivery ratio has been confirmed as the relationship function between the soil erosion and the sediment yield (Dutta 2016). About 80% of the world’s agricultural land suffers from moderate to severe erosion (Speth 1994), and about 75 billion metric tons of soil per year are removed (Pimentel et al. 1995). For example, it is reported that about 1.6 million metric tons per year of sediment is deposited in Chilika Lake by rivers and streams, which is continuously becoming shallow and sandbanks (Bengtsson et al. 2012). Increased water use, increased evapotranspiration, and decreased rainfall result in decreased freshwater volumes and flow rates. This is also considered to exacerbate the deposition of sediment particles by gravity and their consolidation over time (Dodds 2002).

On the other hand, sediments are the source of various pollutants. Sediment pollution as one of the most serious environmental problems in the aquatic ecological system has attracted international attention, both scientific and political, for its potential toxic threat to aquatic organisms and ecosystem (Yang et al. 2016). For example, about 5.3% of the UK environmental polycyclic aromatic hydrocarbon burden is ultimately associated with freshwater sediment (Wild and Jones 1995). More than 99% of aquatic heavy metals are stored in the sediment in various forms (Salomons and Stigliani 2012). In addition, sediments are polluted by eutrophication due to the adsorption of agricultural nutrition like nitrogen and phosphorus (Mateo-Sagasta et al. 2017). Hence, sedimentation management is of critical importance not only for the sustainable development of freshwater resources, but also for the freshwater flora and fauna and the entire water ecosystem.

Sediment dredging is the most commonly used method not only to maintain the sailing depth and flood capacity but also to rapidly reduce the pollution stress (Olsen et al. 2019). For a long-term perspective, proper dredging management has a positive impact on the ecology, water transparency, sediment balance, nutrient budgets, and river morphology downstream of the reservoir (Sumi and Hirose 2009). However, the sediment dredging is still very controversial due to the uncertain negative effects. The agitation and removal of the sediment will dramatically accelerate the secondary release of pollutants in the sediment in a short period of time and seriously change the living environment of benthic microorganism, animal and plant communities (Wang et al. 2014). Environmental window concept has been proposed to protect sensitive biological resources or their habitats from potentially detrimental effects of dredging and disposal operations soon after passage of the National Environmental Policy Act in 1969 (Reine et al. 1998).

The effect of dredging on polluted water bodies is the focus of dredging research and application. Due to the far-reaching effects of dredging on freshwater systems, it is necessary to review and study the relevant literature and make some policy recommendations. This paper systematically reviews the literature on sediment dredging and categorizes its impacts. The impacts of dredging are centered around nutrients, heavy organic matter, and biological communities. The paper also considers short-term and long-term impacts. This paper elucidates the mechanisms underlying changes in the types and levels of various substances in freshwater systems before and after dredging. The paper will contribute to the identification of potential environmental risks associated with bottom dredging.

Research trend of freshwater dredging engineering

Sediment dredging engineering refers to a series of activities such as determining the depth and volume of dredging in an exact area, selecting reasonable dredging equipment, and formulating corresponding secondary pollution prevention programs such as avoiding diffusion and resuspension of fine particles (Yell and Riddell 1995). Therefore, different dredging designs and different dredging equipment have different environmental impacts.

Thematic bibliometrics (a common tool to assess scientific production through mathematical and statistical methods) (Pritchard 1969) was used to identify the growth and hot topic in the field of freshwater sediment dredging. Web of Science (WOS) core collection was chosen as the sources of literature database. The search strategy (TS = “fresh*water” or “river” or “lake” or “reservoir” or “wetland” and “sediment* dredge*”) was used to search the related research from their title, abstract, or keywords, in a time span of 1900–2021. In order to pursue valuable literature, the abstracts of references were browsed to identify and discard repetitive research as well as publications in other fields. A total of 1453 publications were identified as being related to freshwater sediment dredging distributed in 61 Web of Science categories. The distribution of research in the top 10 fields can be seen in Fig. 1a. It is obvious that environmental science studies were dominant, with the contribution of 52.44% of the total selected publication.

Fig. 1
figure 1

a The publication distribution in Web of Science categories. b The growth trend of the publications

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In order to contrast with related research on sediment dredging that not just freshwater, we changed the search strategy to (TS = “sediment* dredge*”) with other conditions unchanged. The growth trend of related research can be seen in Fig. 1b. Research on marine sediment dredging dates to 1937, and research on freshwater sediment dredging dates to 1969. From 1990 the aqueous sediment dredging has attracted international attention and entered a rapid development period. Not only from the perspective of research history but also from the perspective of publication volume, the development of freshwater sediment dredging is relatively weak. Therefore, it is necessary to draw on the relevant experience of marine sediment dredging based on a comprehensive summary of the problems related to freshwater sediment dredging.

Effects of dredging on physicochemical properties of freshwater systems

Hydrological characteristics and basic physicochemical indicators

Due to dredging excavation, large holes are formed at the dredge site in the basically flat-water bottom. Water velocities increased at the leading edge of the hole and accompanied by upstream erosion, caused by the increased gradient of water bottom (Fischer et al. 2012). Whereas, water velocities are decreased within the dredge hole, which allows the suspended sediments to fall out of the water column. Afterwards, the sediment-starved water erodes the tail end of the hole when it leaves the hole with rein creased velocities.

The resuspension of riverbed material due to dredging is a non-negligible negative impact. It can lead to turbidity (Pennekamp et al. 1996; Lu et al. 2019), which regulates light transmission in the water and thus affects photosynthesis in submerged vegetation (Erftemeijer and Lewis III 2006). However, due to the complexity of climatic conditions, the width and depth of the watercourse, and upstream and downstream topography, it is difficult to determine the degree, diffusion, reduction of turbidity and its impact on the entire watershed (Grasso and Le Hir 2018; Vagge et al. 2018). Dredging also affects hydrothermal patterns. This is mainly due to the fact that increased water depths and expanded water volumes result in much slower channel flow rates and reduced vertical mixing (Kaur et al. 2007). Ding et al. (2019) found the temperature continues to decline throughout the whole dredging period, which is not immune to impact the input and decomposition of nutrient and furtherly alter the biological activities (Brönmark and Hansson 2002). In addition, dredging can cause noise and stench.

Physicochemical properties of pH, dissolved oxygen (DO), and oxidation reduction potential (Eh) in undredged (UDR), freshly dredged (FDR), and post-dredged (PDR) sediment–water interfaces are commonly used observation indicators. The mean pH values in FDR and PDR water were lower than UDR (Ding et al. 2018; Saeki et al. 1993). The UDR area has a lower pH from 9.13 to 8.51 than the dredged area under the same climatic conditions (Chen et al. 2018). However, the contradictory result has been reported in the literature that the impacts of hydraulic dredging on surface water pH was negligible (Zhang et al. 2010a; Lewis et al. 2001). It can be speculated that the change of pH value is mainly determined by the content of acid-volatile sulfides in the sediment. The enhanced acidity of the FDR and PDR water is caused by the exposure and oxidation of the acid-volatile sulfides in the original bottom sediments, resulting in the release of a large amount of hydrogen ions (Gambrell et al. 1991; Borma et al. 2003). The most reactive fraction of sulfides in natural sediments is primarily FeS, and its oxidation process is as following (Eq. 1) (Tao et al. 2005).

$$\begin{array}{*{20}l} {FeS_{2} + 14Fe^{3 + } + 8H_{2} O \to 15Fe^{2 + } + 2SO_{4}^{2 - } + 16H^{ + } } \hfill \\ {FeS + 8Fe^{3 + } + 4H_{2} O \, \to \, 9Fe^{2 + } \, + \, SO_{4}^{2 - } + \, 8H^{ + } } \hfill \\ \end{array}$$
(1)

DO in UDR was reported lower than that in FDR and PDR water under the same aeration frequency. It was reported that DO was increased rapidly to saturation level of 8.6 mg l−1 around the fourth day after dredging (Tao et al. 2005). Whereas, different results in other researches have shown that the DO in FDR was the lowest among these three stages (Jing and Li 2016). On one hand, the short-term oxygen consumption of reducing substances in the exposed deep sediments will induce the decrease of DO in the water system (Morgan et al. 2012); on the other hand, the turbidity caused by dredging will also reduce DO through inhibition of photosynthesis by hydrophytes and phytoplankton (Meng et al. 2018). But, from a long-term perspective, dredging has the potential long-term effect of reducing dissolved oxygen depletion capacity (Liu et al. 2006). Therefore, it is recommended that the dredging course should be avoided in the summer to avoid high biological activity that can aggravate the water’s hypoxia (Kaur et al. 2007; Chen et al. 2018). Many pieces of research have shown that DO is a crucial factor influencing the Eh value, and a positive correlation between these two factors was found (De Jonge et al. 2012; Eggleton and Thomas 2004). The enhancement of Eh was likely attributed to a rapid and significant increase in DO in the water (Pourabadehei and Mulligan 2016).

Concentration and speciation of nutrients

Dredging can remove surface sediments, which usually contain a high content of organic matter. Hence, it has been wildly used to address the troublesome issue in eutrophic watersheds (Peimin et al. 2000; Zhang et al. 2010a; Liu et al. 2015a; Holmer et al. 2003). The most typical elements of the limiting nutrients in eutrophic freshwaters are nitrogen and phosphorus (Schindler 1971; Zhang et al. 2013; Tyrrell 1999). Transportation and transformation process of nitrogen and phosphorus in freshwater is shown in Fig. 2. The concentration and speciation of nutrients in the watershed are directly and indirectly affected by dredging engineering, due to the disturbance for photosynthesis, bioaccumulation, high-content surface sediment removal, resuspension, etc. Many studies have strived to explore the nutrients accumulation and transformation in pore water and overlying water to evaluate the short-term and long-term effectiveness of dredging (Wang and Feng 2007; Yenilmez and Aksoy 2013; Recknagel et al. 1995). Yu et al. (2016) have suggested that nitrogen cycling rate in sediments was slowed down by the excavation of sediment which contains a lot of organic matter deposition and microbial community. However, the effectiveness of sediment dredging on eutrophication is still debatable. Contradictory results have been suggested among various studies, and these discrepancies may be attributed to the implications of pollution status of the sediments and dredging technics (Fan et al. 2004; Lohrer and Wetz 2003; Peimin et al. 2000). For example, in the overlying water, the phosphorus concentrations decreased from 0.127 to 0.081 mg/L in April, while nitrogen concentrations increased from 1.84 to 2.38 mg/L in October, and the trends of nitrogen/phosphorus rise or fall are completely different in different seasons (Chen et al. 2018). And the removal of surface sediments did not significantly reduce the concentrations in sediment pore water (Chen et al. 2018). Many studies proved that the nitrogen concentration will increase in the short term after dredging because of disturbance and suspension, but it will decrease in the long term (Jing et al. 2019; Morgan et al. 2012; Zhong and Fan 2007). Reddy et al. (2007) confirmed that the net phosphorus release or retention was linearly related to the rate of phosphorus loading/initial sediment concentration. Even though, removal of the top 30 cm of sediment can remove approximately 65% of total phosphorus sediment storage. When the loading rate was as low as 9.4 mg/m2/year, a linear increase in phosphorus release occurred during the first 156 days after dredging (Reddy et al. 2007). Kleeberg and Kohl (1999) also argued that dredging of surface sediments alone will not reverse eutrophication, unless external loads are also curtailed. Therefore, the reduction of external loads is a key factor for the ultimate effect of dredging on nutrients limitation.

Fig. 2
figure 2

Transportation and transformation of nitrogen and phosphorus in the freshwater system

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In addition, the distribution of nitrogen in different common forms in the freshwater system (\(NH_{4}^{ + }\), \(NO_{3}^{ - }\), \(NO_{2}^{ - }\), and organic N) is also affected by the dredging projects, and this effect is only short-term (Voutsa et al. 2001; Yu et al. 2016). And the denitrification was weaked in the dredged sediment–water systems which declined the attenuation of nitrate and further enhanced the nitrate content. Yu et al. (2016) also discovered that the fluxes of \(NO_{3}^{ - } - N\) from PDR sediments to overlying water significantly increased by 58% and the corresponding fluxes of \(NH_{4}^{ + } - N\) decreased by 78.2% after dredging. However, this result is dramatically related to the external N loadings and the re-sedimentation of suspended particles. Liu et al. (2019) found that the labile \(NH_{4}^{ + } - N\) concentration increased from 128.24 mg g−1 to 296.75 mg kg−1 in the surface sediment after dredging and the ammoniacal nitrogen concentrations in pore water of two dredged groups with external loadings were higher than those of UDR groups. Rise of NH4+ concentration may be attributed to the two reasons. Ammonia is predominantly present as NH4+ loosely bound onto the circumneutral sediments as an exchangeable fraction, which is easily resease into waters disturbed by dredging. And organic nitrogen compounds and oxides release NH4+ by dissolving oxygen (Eq. 2) (Choppala et al. 2018).

$$(CH_{2} O)_{a} (NH_{4}^{ + } )_{b} + aO_{2} \begin{array}{*{20}c} \to \\ \end{array} aCO_{2} + bNH_{4}^{ + } + aH_{2} O$$
(2)

The effects on fractions of phosphorus in the water–sediment system were also well studied similar to the researches on the accumulation and transformation of nitrogen during dredging engineering. Wen et al. (2020) reported that after dredging, the content and proportion of mobile phosphorus (the sum of NH4Cl-P, Fe–P and Org-P) in the surface sediments were significantly reduced. Attribute to the improvement of DO, the Org-P in the surface sediment of FDR release soluble phosphorus rapidly (Eq. 3), while the \(Fe\left( {II} \right)\) monosulfides rapidly oxidized to \(Fe\left( {III} \right)\) oxide minerals possessed stronger retention capacity for \(PO_{4}^{3 - }\), which in turn decreased the phosphorus concentration in the overlying water. Moreover, the mineralization of Organic-P is easily adsorbed and combined with calcium ions and converted into Ca-P and Residual-P resulting in the transformation of phosphorus from active to inert fractions. However, the proportion of labile phosphorus (diffusive gradients in thin films and soluble reactive) in pore water of the FRD group was enhanced due to the release of phosphorus from the resuspended matter. But the external loading identically diminished the effectiveness of dredging.

$$(CH_{2} O)_{a} (H_{3} PO_{4} )_{c} + aO_{2} \begin{array}{*{20}c} \to \\ \end{array} aCO_{2} + cH_{3} PO_{4} + aH_{2} O$$
(3)

Activation and release of heavy metals and organic pollutants

Sediment dredging is one of the most used means to control endogenous pollution. It removes contaminated river sediments and reduces the possible flux of heavy metals. (Ding et al. 2015a). But some scholars believe that thin layer dredging could temporarily reduce total sediment metal concentrations but not heavy metal bioavailability. Heavy metals do not degrade during dredging but transform between soluble and insoluble forms (Peng et al. 2009; Akcil et al. 2015). Continuous extraction, vitrification, and thermal and biological treatments are often used for the assessment and treatment of heavy metals in sediments (Mulligan et al. 2001; Meers et al. 2005; Kim et al. 2011). Studies have shown that the sediment dredging operation has induced the resuspension of sediment particles. These contain heavy metals deposited by sulfate or adsorbed on organic matter. The root cause of heavy metal pollution that cannot be alleviated after dredging may be the resuspension of the contaminated sediments. Similarly, Liu et al. (2016a, b) have found that heavy metal concentrations (As, Cd, Cr, Cu, Ni, Pb, and Zn) in the FDR surface sediment increased to UDR levels with the influence of metal-adsorbed suspended particulate matter, especially Zn and Cd increased 482.98% and 261.07%. Although the total metal content in the dredged sediment is greatly reduced, the increased heavy metal concentrations were mostly in the relatively bioavailable non-residual fractions (Yu et al. 2019). Wasserman et al. (2013) even straightforwardly points out that the dredging of contaminated sediments is a harmful activity for the environment largely due to the contaminant’s resuspension and bio-uptake.

The release mechanism probably depends on the characteristics of metal-adsorbed suspended particulate matter including fine grain size and the pollutant contents (Liu et al. 2016a, b). In the simulation study conducted in Lake Taihu, Yu et al. (2019) figured out that the release of heavy metals after dredging was attributed to the migration of metal sulfides in the deep sediments that are anoxic. Dredging introduces oxidized water, which increases oxidation. This causes sulfide-bound metals to be dissolved by sulfate-reducing bacteria (SRB) into available volatile sulfides (AVS) (Fig. 3). However, the concentrations of trace elements in the water column were not affected by dredging, except for Cu and Zn, which were attributed to the pH neutrality of the sediment and the formation of metal sulfides that were strongly retained in the solid phase (Choppala et al. 2018). In addition, the impact of dredging on the release of heavy metals is also strongly related to time and season. Chen et al. (2019a, b) pointed out that dredging in winter had a positive impact on the release of Co, Zn, and Ni, with the largest increase in soluble Zn and Co in the overlying water in January after dredging, an increase of 166% and 69%, respectively. Sun et al. (2019) also confirmed that dredging effectively remediated metalloid contamination (arsenic, selenium, and antimony) in sediments only in April, July, and/or January, but negligible/negative effects were seen in October. The impact of time and season on the release of heavy metals might be caused by the degradation of algae in autumn and winter, because the decomposition of algae might reduce oxygen saturation, thereby accelerating the release of metals in sediments (Chen et al. 2019a, b; Yang et al. 2020). Furthermore, after dredging, the influx of high-level metal adsorbed suspended particulate matter from exogenous polluted rivers may also adversely affect the sediment–water interface, thereby increasing the risk of heavy metal pollution in the water environment (Liu et al. 2016a).

Fig. 3
figure 3

The effects of dredging on the migration and transformation of heavy metals (Yu et al. 2019)

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In addition, owing to the organic pollutants such as polycyclic aromatic hydrocarbons are hydrophobic, they are more likely to settle and accumulate in sediments (Kafilzadeh 2015), which means that like heavy metals, during the dredging process, the resuspension of sediment, the release of pore water in the bottom sediment, and the adsorption of suspended and resuspended substances will all cause the re-release and diffusion of organic pollutants (Qi et al. 2011). Many studies have confirmed this. Ruocco et al. (2020) believed that dredging might result in the resuspension of sediments and the release of organic pollutants such as polycyclic aromatic hydrocarbons, polychlorinated biphenyls, and organotin compounds, allowing them to diffuse into the water column, which may affect the water environment. Cutroneo et al. (2015) confirmed that dredging was the key factor of rising the mobilization of polycyclic aromatic hydrocarbons concentrations in the water column. This is because dredging removes the surface sediment layer, allowing the organic matter in the deeper sediment layer to be released into the overlying water column (Zhong et al. 2010). Two to three rings polycyclic aromatic hydrocarbons (PAHs) (maximum value 0.105 µg/L) presented the largest concentration in dredging operations and then decreased rapidly because it was more soluble in water, more biodegradable and easier to evaporate, while four rings PAHs predominated at the phase of PDR with the average concentration of 0.057 µg/L (Vagge et al. 2018). The variation of PAHs in the overlaying water can be seen in Fig. 4.

Fig. 4
figure 4

Concentrations of PAHs according to the number of rings (2–3 rings, 4 rings, 5–6 rings, and HMW (high molecular weight)) during phases UDR (undredged), FDR (fresh dredged), and PDR (post-dredged). Data quoted from (Vagge et al. 2018)

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Disturbance to biotic communities

The ecological effects of substrate dredging are also reflected in changes in benthic fauna, phytoplankton species and abundance, and biomass and community structure in the water column. Aquatic biotas govern the cyclic processes of nutrient, energy and organic matter and maintain the dynamic balance of water system. Dredging has the potential to impact aquatic biota through direct entrainment into the dredges and indirectly altering the water and sediment habitats, respectively (Manap and Voulvoulis 2014). For instance, the effects on certain aquatic organisms can be catastrophic due to the increased water turbidity caused by dredging, which interferes with the predation and respiration of some aquatic fauna (Aldridge 2000) and the photosynthesis of submerged vegetation and algae (Freedman et al. 2013). Dredging can also cause macrophyte beds and invertebrates habitats to be destroyed and fish lost, as sediments are removed (Freedman et al. 2013). Additionally, dredging disturbance could significantly change the composition and structure of sediment communities. Given the ecological relevance, changes in overall community trophic and functional structure can eventually affect the entire biological chain (Coates et al. 2015). Such impacts can take decades to recover (Haynes and Makarewicz 1982; Boyd et al. 2005; Waye-Barker et al. 2015), and even when abundance, biomass, and species numbers recover, the original composition, structure, and ecosystem may have been disrupted (Barrio-Froján et al. 2008, Barrio-Froján et al. 2011), and such impacts are long-term (Szymelfenig et al. 2006). For instance, the study of Zhang et al. (2017) indicated that dredging could stimulate aerobic populations but repression of anaerobic groups, resulting from increased sediment dissolved oxygen and oxidation–reduction potential, and one year after the dredging, the microbe and macrobenthos communities had reached alternative state instead of returning to their original state. Therefore, the effect of dredging on biotic communities is a very serious issue that deserves attention.

Fishes and invertebrates

Fishes

Freshwater fisheries are globally important ecosystem services with subsistence, commercial and recreational value. Fish are part of a complex ecosystem of interconnected habitats and organisms; hence, perturbation from sediment dredging threats fish populations through pathways (Fig. 5). Dredging can lead to the loss of fish habitat (Erftemeijer and Lewis 2006; Erftemeijer et al. 2012). Habitat degradation and loss can have a significant impact on fish communities (Galzin 1981). Suspended sediments from dredging reduce fish response distances and visual acuity negatively impacting pelagic fish foraging (Sweka and Hartman 2003; Zamor and Grossman 2007). In addition, suspended sediments can damage gill structure and tissue with physiological effects on exposed fish (Hess et al. 2015). Freedman et al. (2013) quantified the differences in fish assemblages sampled by benthic trawls among dredged and undredged sites in the Allegheny River, Pennsylvania, USA by ecological metrics and stable isotope analysis. The results suggested that the habitat loss caused by gravel dredging reduced the benthic fish abundance and diversity due to the lack of suitable spawning habitat or reduced foraging efficiency. Hayer and Irwin (2008) reported that instream gravel dredging contributed to 38% of the species variation in detection probabilities of 87 Piedmont and Coastal Plain fishes collected in four Alabama streams of the Mobile River drainage, and fish species that prefer riffle habitat and coarse substrate were lower in dredged areas, whereas species that preferred fine substrate were more abundant in dredged areas. Conversely, dredging operations may provide multiple habitat types; thus, the response of fishes to dredging is unclear (Rempel and Church 2009). The high-velocity habitat directly upstream of the nickpoint may be beneficial to lotic fish species, while lentic species prefer to be present in dredge holes (Paukert et al. 2008). Additionally, the majority of these studies were conducted in small scale rivers and short-term effects. Kjelland et al. (2015) reviewed those numerous studies have demonstrated the effects of suspended sediments during dredging on exposure and mortality rates of affected fish, deposited eggs, or larvae, and the long-term effects on epigenetic changes should be further conducted.

Fig. 5
figure 5

Potential effects of dredging on fish and fish habitat generalized by Fishers and Oceans Canada (referenced from (Ward-Campbell and Valere 2018))

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Benthic fauna

The survival of most benthic communities requires a suitable habitat, such as a certain amount of organic sediment, while excessive dredging could pose a large-scale anthropogenic disturbance to it (Zhang et al. 2014). In particular, the macrobenthos (> 0.5 mm) are used as important indicators of dredging disturbance because they inhabit different substratas exhibit varied behaviors and feeding patterns in response to their different functional needs. And macrobenthos play crucial role in the nutrient recycling, secondary production and pollutant metabolism, dispersion, and burial (Rehitha et al. 2017). Dredging would have some short-term effects on the benthic fauna. For example, benthic animals were removed directly (Jing et al. 2019), resulting in a decrease in abundance, which was directly related to dredging disturbance events (Lewis et al. 2001; Freedman et al. 2013; Rehitha et al. 2017; Bettoso et al. 2020). Meng et al. (2018) proved that there was an alteration in the composition and a reduction in biodiversity of the benthic fauna in the disturbed area, while at the same time, an increase of them in the nearby region, which further shows that benthic communities are more willing to avoid unfavorable living environments. Dredging also completely destroyed the habitat, leading to long-term disruption and alteration in the benthic community, which might need to take several years to recover, with the possibility of never recovering (Ceia et al. 2013), profoundly affecting aquatic ecosystems (Zhang et al. 2014). There are studies having demonstrated that benthic communities were very sensitive to the disturbance of dredging (Zou et al. 2019). For example, many macroinvertebrates live in the sediment for a long time, and the ability to avoid the negative effects of dredging is very weak due to the regular staying still. Moreover, habitat degradation caused by dredging could impede the recolonization of macroinvertebrates, therefore, there would be a considerable decline of the number of macroinvertebrates after dredging (Zou et al. 2019). In addition, for the molluscs, dwelling in the bottom of the water body lead to their preference for a stable substratum, so the unstable substratum caused by dredging would contribute to the decreased density of molluscs (Rehitha et al. 2017).

The adverse effects of dredging on benthic organisms are a major concern for many scholars. Ceia et al. (2013) pointed out that an obvious consequence of dredging activity was the short-term decrease in species density and biomass. Although the number, diversity, and ecological index of the microbenthic communities would gradually return to pre-dredging situation after 2 months of dredging, the macrofauna composition and structure were still dissimilar, and the new assemblage might not have the same ecological functions. This may be attributed to the fact that dredging provides a window of opportunity for new species, enabling them to settle in physically disturbed areas and benefiting from the imbalance of the initial population (Piló et al. 2019). It can be seen that the disturbance caused by dredging contributed to the reduction of sensitive species, with replacing them with tolerant species (Licursi and Gómez 2009; Netto et al. 2012). During the dredging of South Lake Dongting in China, Meng et al. (2020) found that the structure of microbenthic taxonomic and functional assemblages would be strongly affected by the disturbance process, resulting in the simplification of the macroinvertebrate functional structure in the dredging area, which can be proved by the loss of certain feature categories. Therefore, dredging has profound negative effects on benthic communities, and necessary measures should be taken to reduce these effects during the dredging process, such as avoiding breeding period and controlling dredging depth to reduce the destruction of biological habitats.

Zooplankton

In water ecosystem, zooplankton is an important contributor to energy transfer from primary producers to higher nutrient levels (Nandy and Mandal 2020; Abdullah et al. 2017; Degerman et al. 2018), and zooplankton species, abundance, and biomass are often used as important indicators to assess water quality due to its high density, short life span, natural drift and high sensitivity to environmental fluctuations (Gorokhova et al. 2016; Ptacnik et al. 2009). The trophic state of the water column is a key factor affecting the zooplankton community and changes in environmental conditions caused by dredging may lead to changes in the zooplankton community (Karr 1981; Sládecek 1983). Zooplankton abundance and dominant populations can also be a useful indicator of the impact of human activities on the ecology of a water body, and changes in water environment factors can also affect zooplankton community structure (Suikkanen et al. 2007). Zooplankton abundances are higher in the UDR than in PDR areas reduced by large-scale dredging (Lurling et al. 2017). For example, Zhang et al. (2010b) proved that the decrease in the abundance of rotifers was a response to dredging. This appears to be attributable to a reduction in food supply. Because dredging would change the resource utilization and nutritional pathways within the food web (Freedman et al. 2013), which is closely related to the zooplankton community composition. In addition, the polluted suspended sediment caused by dredging contains heavy metals, organic pollutants, etc., which are toxic to plankton including copepods, and it is also one of the reasons for the negative impact of zooplankton (Sew et al. 2018).

Aquatic plants

As a major producer of aquatic ecosystems (Wu et al. 2019), phytoplankton plays an important role in the water environment, and it is considered to be an effective indicator of water ecological changes (Cabrita 2014) due to its quick response ability (Cabrita et al. 2013). Take phytoplankton with simple structure like microalgae as an example, it is considered to be an universal and reliable indicator of the watershed system ecology for their wide distribution, quick generation time, distinct community structure, and specific response to habitats conditions (Thomson and Manoylov 2019). Hence, sediment dredging projects could have an inevitable negative impact on phytoplankton by disturbing water bodies and sediments. For example, the resuspension of contaminated sediments caused by dredging could result in observable toxicity (Nayar et al. 2004), which changed the phytoplankton community structure (Cabrita 2014), particularly influencing the small phytoplankton cells (Lafabrie et al. 2013).

There are studies that have pointed out that, compared to the UDR conditions, the concentration of Cu, Cd, Hg, and Pb in the phytoplankton cells in the PDR environment was significantly higher (Cabrita et al. 2014), which would significantly inhibit the growth rate of the cells, thereby limiting the growth of phytoplankton (Cabrita et al. 2013). This shows that the metal elements released during the dredging process did affect the physiology of phytoplankton. Moreover, Nayar et al. (2004) had also proved that bioavailable heavy metals in the resuspended sediments after dredging might have a large-scale negative impact on the biota, especially on phytoplankton. In addition, there are also reports on the impact of dredging on phytoplankton biomass and abundance. Thomson and Manoylov (2019) believed that dredging could result in the observed diatom species richness to decrease, and the results of the study in Chen et al. (2015) showed that due to frequent anthropogenic activities such as dredging, the species of phytoplankton decreased significantly, especially the diatoms, decreased by 70%, accounting for most of the decreased species of phytoplankton in the Changjiang Estuary. Furthermore, Jing et al. (2019) proved that during the dredging of the South Lake, the phytoplankton biomass decreased with time as the dredging project progressed, and the phytoplankton biomass reached the lowest value when the dredging project was completed.

Another important pathway of the impact is that suspended particles caused by dredging are usually finer than natural coarse sediments. It can seriously decrease the light availability of benthic algae, inhibiting the biomass of these microphytes (Robinson et al. 2005), and these suspended particles might take longer to settle out of the water column (Cunning et al. 2019). The benthic diatoms could be adversely affected as well. In the process of dredging, the increase of turbidity and suspended solids would reduce the penetration of light, causing the density of benthic diatoms to drop immediately after dredging (Licursi and Gómez 2009).

On the contrary, some studies have proved that dredging would not lead to a decrease in phytoplankton biomass, especially in the short term. Lafabrie et al. (2013) believed that brief sediment resuspension events caused by dredging might be beneficial to phytoplankton communities, because although the contaminated sediment would bring about the release of toxic chemicals into the water, it could also strongly stimulate phytoplankton growth within a few hours. However, they also pointed out that the beneficial effects were only short-term, and in the long-term, dredging is still harmful to this important autotrophic component due to toxicity.

Microorganisms

The microbial communities play important roles in the function of lotic ecosystems and become key players in nutrient cycling by the function of decomposing organic matter and degrading various pollutants (Liao et al. 2020). In the water environment, they can affect the circulation of nutrients such as nitrogen, phosphorus and carbon through metabolic activities, further affecting the biogeochemical cycles in sediments, and in turn affect the microbial community themselves (Liu and Yang 2020). Environmental genomics also regards microbial communities as potential biological indicators of environmental conditions (Baniulyte et al. 2008), because the composition and functional characteristics of the microbial community would change following a drastic habitat disturbance (Sui et al. 2020). For instance, in the process of dredging, the ecologically-meaningful planktonic microbial could quickly respond to the resuspension of sediments caused by dredging disturbances, such as a rapid decrease in abundance; therefore, it can be used as a useful tool to evaluate the impact of dredging on the ecosystem (Layglon et al. 2020). Liao et al. (2020) believed that sediment dredging might immediately lead to a significant loss of microbial diversity, because of the greater vulnerability of water microbial communities under intensive watershed disturbances. Studies have also proved that some microorganisms attach onto the surface of suspended sediments for a long time by secreting sticky extracellular polymeric substance, and the number was even 4 to 7 times that of free-floating ones (Lind and Dávalos-Lind 1991). Therefore, the dredging to remove sediments would inevitably cause the loss of microorganisms.

However, for the total microbial biomass, there might be no significant change after dredging (Baniulyte et al. 2008), since the pollutants and nutrients in the surrounding environment were easily absorbed by the suspended sediments in the water body, which contributed to the growth of microbes and improved the survival rate of them (Ramalingam and Chandra 2018). Nevertheless, this could not avoid the negative effects of dredging, owing to the significant changes in the structure and diversity of microbial communities. The study of Wan et al. (2020) indicated that, after dredging, although the plankton bacterial community might display stronger resistance to environmental changes in the short term on account of the insignificant correlation between the physicochemical factor and bacterioplankton community function, in the long run, dredging decreased the diversity and function of the bacterioplankton community, and led to nutrient deficiency, further causing a decline in the regulation of the bacterioplankton community. This phenomenon might imply that microbial community did not have stronger anti-interference ability, and there was a high possibility that it could not return to the initial stable state after dredging. In addition, the dredged cores were extremely susceptible to disturbance, which would cause the microbial community on the sediment surface before dredging to be unable to recover for a long time in the future (Qian et al. 2012). Zhang et al. (2017) also deemed that, after dredging, the micro eukaryotic and as bacterial communities were different from their original state. And one of the reasons that caused community change but the possible exception of biomass might be community succession, where sensitive species were replaced by more tolerant species, which was also one of the responses of microorganisms to the disturbance of dredging (Zhang et al. 2017). In general, the impact of dredging on microorganisms is relatively large, but not many studies have been reported so far. Therefore, the microbial response to dredging interference should be further explored in the future.

Problems and suggestions

As a common engineering method to control internal load of nutrients and pollutants, sediment dredging has been used in many freshwaters. Although positive effects have been achieved, there are still the following problems, including (1) dredging can expose deep sediments to the overlying water, change the physicochemical environment, and release and activate various elements such as nitrogen, phosphorus, heavy metals, and organic pollutants at the sediment–water interface. (2) Based on the above negative impacts, dredging will further irreversibly affect the biomass and community structure of aquatic organisms in freshwater. For this reason, prior to the dredging operation, a preliminary assessment should be made to reasonably predict the dredging parameters including the dredging time, depth, and frequency, so as to find an optimized dredging method to minimize the negative effects of dredging. Below are some suggestions and directions for future research.

  1. 1

    Determine a reasonable dredging depth

    Generally speaking, the concentration of contaminations in sediments gradually decreases as the depth increases. Even though the dredging depth needs to be deep enough to effectively remove contaminations from sediments. However, when it exceeds a certain limit, the deeper the dredging depth is, the more serious the damage to the ecosystem will be. Conversely, if the dredging depth is too shallow, it will not only fail to remove the polluted sediment, but also aggravate the contaminant release. The study of Liu et al. (2015) proved that in Lake Taihu, China, dredging at different depths had significantly different effects on preventing black flowers. And the control effect is the best when the depth was 22.5 cm, the result might attribute to the low levels of acid-volatile sulfide (the most important limiting factor for the occurrence of black flowers in the overlying water) in the sediments after dredging, which reduced the release of hydrogen sulfide to the overlying water. In terms of the control of heavy metal pollution, the dissolved organic matter released from sediments with a depth of 20–30 cm was proved to possess a higher metal binding capacity than those from surface layer sediment. The biological toxicity related to the free ion activity after dredging was reduced (Xu et al. 2016). And Ding et al. (2015) calculated the dredging depth of ten heavy metal-polluted rivers in Pinghu, Zhejiang Province, based on the critical-risk-depth method, and the result suggested that the dredging depth of the ten rivers should be within the range of 35 to 100 cm in order to minimize ecological risks. The depth of dredging can affect different freshwater lakes or areas differently.

    Therefore, field investigations and experimental studies are essential before dredging. Specifically, we need to understand the physical characteristics of the sediments at the dredging site, the current and wave conditions, and the vertical distribution and concentration of the contaminants. Then, we need to analyze and estimate the distribution of target contaminants in the sediment profile and their ecological risks. Finally, based on the above information, we determine the appropriate dredging depth for optimal dredging.

  2. 2

    Choose the right dredging season

    In order to reduce pollution, it is significant to choose the season of dredging reasonably. For example, shortly after dredging, the concentration of NH4+-N in pore water will increase, and this release rate will be highly exacerbated in summer (Liu et al. 2016a, b). Based on this reason, the dredging should be carried out in the lower temperature season. And for benthic organisms, winter is the best option because it avoids the breeding period and the growth period of larvae. Dredging work in winter could minimize the impact on aquatic organisms. However, the research of Chen et al. (2019a, b) on the dredging effect of Lake Taihu proved that the capability of dredging to control heavy metal pollution fluctuated over seasons, being positive in spring and summer and negative in winter. Chen et al. (2019a, b) also believed that in April, July, and October, dredging effectively reduced the soluble and unstable heavy metals in the sediments of Lake Taihu and hindered their leaching into overlying water, but the effect was completely opposite in January. Thus, considering the control of heavy metal pollution, the dredging should be carried out in spring and summer instead of winter.

    Therefore, the determination of the dredging time cannot be generalized. It is necessary to make a reasonable prediction based on the characteristics of the specific dredging area, comprehensively consider various factors, thereby to achieve a satisfactory dredging effect.

  3. 3

    Determine the appropriate dredging frequency

    In addition to the dredging depth and time, the determination of the appropriate dredging frequency is also an important factor to maintain good dredging results. Increasing dredging frequency can alleviate the pollutant aggravation caused by external input, while too frequent disturbance will not only increase the dredging cost, but also irreversibly disrupt the aquatic biological systems. Chen et al. (2019a, b) concluded that in Lake Taihu, the ability of dredging to control endogenous nutrients diminished over time. They suggested a 5-year cycle to consolidate the positive effects. In Chaohu Lake, more frequent dredging operations are needed to reduce the fluxes of P and N in the sediments due to continuous external inputs of suspended particulate matter and residues (Liu et al. 2019).

    Therefore, we recommend that a reasonable dredging frequency schedule be developed based on sediment characteristics and deposition rates. Different sediments behave differently in terms of deposition and resuspension rates. Evaluating these characteristics can help determine appropriate dredging techniques and equipment. For sedimentation rates it helps to gain insight into the dynamics of sediment transport in the water column.

  4. 4

    Combine the dredging process with protective measures

    The negative effects of dredging are unavoidable, but measures could be taken to minimize them while dredging. External measures can reduce the environmental risks associated with sediment dredging. For example, dredging could be combined with other restoration measures. For example, we can restore ecosystems by replanting underwater macrophytes. These plants not only stabilize sediments but also improve water quality. We can employ specialized techniques to retain and dispose of dredged material. It is feasible to utilize constructed wetlands or sedimentation ponds to capture and dispose of nutrient-laden sediments. We can also adjust the method and timing of dredging to minimize disturbance to sensitive habitats and the breeding season. Finally, we can consider recycling clean sediments. Use for beneficial purposes such as land reclamation or construction is also a sustainable option.

Conclusions

By systematically reviewing and analyzing the literature, it is found that since 1990 the aqueous sediment dredging has attracted international attention and entered a rapid development period, but the research on freshwater sediment dredging is relatively weak. Importantly, dredging of freshwater sediments has inevitable impacts on the environment, including secondary releases of nutrients and pollutants to the overlying water system and disturbance and toxicity to aquatic organisms by different pathways during dredging process. However, due to the different hydrogeological conditions of the water body and the different types and concentrations of pollutants in the sediment, the effect and the environmental impact are quite different, and there is no universally applicable law to be found. In general, the effects of dredging on the physicochemical properties, nutrient release, organic matter concentrations and aquatic communities of a water body can be divided into short-term and long-term effects. Overall if dredging is carried out properly it will have a beneficial impact on the freshwater system. The season, location, depth, and equipment of the dredging will also have an impact on the effectiveness of the dredging. In addition, although the utilization of dredged bottom has increased a lot in the world in recent years, the overall utilization is far from adequate. There are no more effective methods to treat contaminants in dredged sediment. These still need to be further explored.