| Citation: | Xiancang Wu, Teng Ma, Yanxin Wang. Surface Water and Groundwater Interactions in Wetlands. Journal of Earth Science, 2020, 31(5): 1016-1028. doi: 10.1007/s12583-020-1333-7
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Wetlands are one of the most important ecosystems in the world, and their total area accounts for 6% of the total land area of Earth (Desta et al., 2012). The type of wetland includes coastal, river, stream, lake, swamp and constructed wetland. Wetlands are helpful in reducing floods, recharging groundwater or augmenting low flow of rivers and influencing regional microclimates, and therefore are very beneficial for reducing natural disasters, such as damage from heavy rain and drought (Acreman and Holden, 2013; Zedler and Kercher, 2005). They can also degrade toxins and pollutants, purify water and reduce environmental pollution (Phillips et al., 2015; Mitsch et al., 2012). Wetland ecosystems are the interactive zone between aquatic and terrestrial ecosystems and are home to many rare species of flora and fauna (Hu et al., 2017). However, in recent decades, wetlands have faced shrinkage of their area, reductions in biodiversity, increasing soil and water pollution and degradation of their functionality (Wu et al., 2017; Davidson, 2014; Smolders et al., 2009).
As an essential part of wetland hydrology, groundwater is important in water and substance balance of wetland ecosystems through surface water and groundwater (SW-GW) interactions, thus plays an essential role in supporting the stability of wetland ecosystems (Havril et al., 2018). SW-GW interactions occur by subsurface lateral flow through the unsaturated soil and by infil- tration into or exfiltration from the saturated zones (Sophocleous, 2002). In wetlands, SW-GW interactions have an impact on the transformation or removal of substances, such as major ions (Lagomasino et al., 2015), nutrients (Du et al., 2017b; Zhou N Q et al., 2014) and heavy metals (Boyer et al., 2018). In arid and semi-arid areas, groundwater may be the most important, possibly the only, source of water for wetlands. In these areas, SW-GW interactions become the dominant factor controlling the formation, development and even extinction of wetlands (Crosbie et al., 2009). Groundwater also has a relatively stable temperature and can regulate the temperature of wetlands and thus provide a more stable temperature for benthic organisms (House et al., 2015). This buffering effect on temperature also influences the hydrogeochemical and biogeochemical reactions in wetlands (Boulton et al., 2008).
Therefore, the evolution of wetlands can be expected because of the changes resulting from SW-GW interactions, and it is necessary to explore the evolution processes and mechanisms from the perspective of hydrogeology (Orellana et al., 2012). The importance of groundwater for maintaining the health, structure and function of wetland ecosystems has received increasing attention. However, in wetlands, SW-GW interactions are more complex, heterogeneous and variable because of the fine-grained and low-permeability sediments and the existence of plants roots (Su et al., 2016; Jolly et al., 2008), so that previous wetland research often neglected or simplified the description of groundwater hydrological processes, and thorough research into SW-GW interactions in wetlands has not been conducted (Wang W et al., 2018). Moreover, with more and more serious global climate change and human activities, such as rising temperature, pollution, drainage and eutrophication, the hydrological conditions in wetlands ecosystems has been deeply changed resulting in the alteration of SW-GW interactions and high vulnerability of wetlands. Hence, it is important to thoroughly learn the SW-GW interactions in wetlands in order to ensure the sustainable development of the wetland ecosystems, and to restore the wetlands which were damaged. Even though researchers have reviewed the SW-GW interactions in wetlands (Fan et al., 2012; Jolly et al., 2008), they failed to reveal the role of these interactions in wetland ecological functions and the development of the approach to identify these interactions.
After carrying out a systematic review of recent literatures, we aimed to examine the characteristics of SW-GW interactions in wetlands and the way to identify these characteristics. Further, we discussed the importance of SW-GW interactions in wetland ecological functions, and showed that climate changes and human activities has threaten the SW-GW interactions and the wetland ecosystems stability. However, there are different degrees of SW-GW interactions, it is inappropriate to divide the interaction into close or weak relationship. Because there is no clear definition between the two, and it has rarely been discussed in previous literatures. So in this paper, we did not highlight the different degrees of SW-GW interactions. The goal of this review is providing a reference to understand the water and material cycling in wetlands and to protect and restore wetlands under climate change and anthropogenic impacts, and to provoke the thorough research in SW-GW interactions in wetlands for better protection regime.
By summarizing previous research results, wetland SW-GW interaction modes can be divided into four types (Fig. 1) (Jolly et al., 2008): (a) connected losing wetlands, where the aquifer is directly connected, and the surface water level is higher than surrounding groundwater, so the surface water becomes the recharge source of surrounding groundwater; (b) disconnected losing wetlands are similar to (a) except that there is a disconnected unsaturated interval between the surface water and the groundwater surface, and surface water vertically recharges the groundwater, which is more common in a seasonal wetland system; (c) flow-through wetlands exhibit hydraulic gradients with a consistent direction of groundwater flow, resulting in surface water receiving groundwater recharge upstream and draining to groundwater downstream, and the groundwater flows through the entire wetland; (d) gaining wetlands, where surrounding groundwater recharge surface water.
However, it is critical to know that individual wetlands may temporarily transfer from one type to another, relying on time-variable dynamics of surface and groundwater within the wetlands. The dynamics of SW-GW interactions of wetlands are strongly controlled by the relative groundwater and surface water levels and these can vary significantly over the short term. Over the long term, changes in GW-SW interactions will occur when there are changes in the water levels driven by factors such as climate change (Sophocleous, 2002), modifications to the management of the uplands and/or riparian zone (i.e., land use change such as clearing of native vegetation for dryland agriculture, irrigation, forestry, urban development, etc.), and changes in the flow regimes of the surface water due to regulation, channelization, upstream water abstractions, etc.. A landscape perspective would consider that channel form, alluvial sediments lateral connectivity and slope, and differential permeability associated with paleochannels and floodplain water bodies (such as ponds, backwater channels and cutoffs) make a very significant difference (Stubbington et al., 2009). At the reach scale, down and upwelling of groundwater may be controlled by discontinuities, such as obstacles, changes in flow direction and pool-and-riffle sequences, which affect the range of vertical hydraulic gradients and hydraulic conductivity (Bertrand et al., 2011).
The zone of mixing between groundwater and surface water is the hyporheic zone (HZ) (Sophocleous, 2002). There is currently no single, inclusive definition of a HZ because of the various perspectives understood by different disciplines. Triska et al. (1989) used geochemical methods in basing a division of the saturated zone containing ≥10% and ≤98% of surface water into a HZ (Fig. 2a), and this is a widely accepted definition by hydrogeologists and is also applied in this paper. They also reported that there were significant physical-chemical gradients in a HZ, including dissolved nutrients, such as nitrates and ammonia, dissolved organic carbon (DOC) and dissolved gases (e.g., carbon dioxide, oxygen and methane). The scope of HZ is broad, with spatial scales ranging from millimeters or centimeters, beneath small bed forms, to longer flow paths of tens, hundreds, or even thousands of meters (Boano et al., 2014) and is dominated by the hydraulic conditions and permeability of sediments.The structure of the wetland HZ has its own unique characteristics that differ from other types of surface-to-ground water HZ. Firstly, owing to the low water flow velocity in wetlands, it is easier for fine particulate matter to be deposited on the wetland bed, forming a layer of fine-grained sediments (Fig. 2a). The sediments usually have relatively low permeability resulting in a decrease in vertical permeability, which further weakens the SW-GW interaction and limits the scope of the HZ. Su et al. (2016) studied the HZ of Dakebo Lake in the Ordos Basin, China and reported that, owing to denser sediment medium and weaker hydrodynamics in the centre of the lake, the thickness of the HZ in the centre of the lake was mainly between 0 and 1.2 m, while that on both sides of the lake was between 2.5 and 4.5 m. The existence of less permeable sediments also led to a debate on whether there is a vertical hydraulic connection between surface water and groundwater. For example, in peatland, previous studies have shown that there is only transverse flow (Price, 1996), and, even if obvious vertical hydraulic gradients are observed, groundwater flow in deep peat is always neglected (Gafni and Brooks, 1990). However, the evidence for a vertical hydraulic gradient from chemical characteristics (Wu et al., 2017) and lithium isotopes (Négrel et al., 2010) supports the view that there is vertical flow in peatland. Therefore, the spatial variability and heterogeneity of wetland HZs are more complex because of the influence of the bottom layer, which has low permeability.
Secondly, there is a considerable amount of vegetation, both supergene and submerged, in the wetlands (Fig. 2a). Vegetation mainly affects the wetland HZ in four ways: (1) by reducing the fluidity of surface water, which is more conducive to the deposition of suspended matter; (2) by plant evapotranspiration, which changes the hydrodynamic conditions; (3) by respiration and nitrogen fixation of the roots, which changes the chemical conditions, and (4) by the existence of plant roots, which changes the characteristics of the porous media and may result in preferential flowing path and priority flow. Studies have also shown that wetland vegetation has significant effects on the groundwater-soil-surface water flux and the exchange of carbon, nitrogen, phosphorus, dissolved oxygen, microbial populations and energy (Zhang et al., 2019; Eamus et al., 2015; Wang et al., 2015). The seasonal water extraction depth and root activity disturb the HZ structure (Jolly et al., 2008). The existence of wetland vegetation makes the physical structure and chemical environment of a wetland HZ more complex, heterogeneous and variable.
There are physical, chemical and biological processes in wetland HZs because of significant physical-chemical gradients (Fig. 2b).
The hydrodynamic gradients result in water exchange between surface water and groundwater which is the dominant factor in biogeochemical and ecological hydrology. The deposition of fine particulate matter causes the clogging of porous media, leading to a decrease in porosity and permeability of the sediment. The clogging process includes clogging of suspended matter due to particulate matter filtration-deposition, gas phase clogging caused by bubble filling and compaction clogging caused by fouling of the leaching layer (Nivala et al., 2012). Physical clogging will lead to slower water flow velocity and a prolongation of residence time, which causes chemical clogging, such as ion exchange/complex, adsorption/desorption and chemical precipitation/crystallization, and biological clogging (i.e., rapid propagation of microorganisms leading to the attachment or accumulation of organisms and metabolites on the surface of the medium particles) (Vymazal, 2018). The porosity and permeability of the sediment changes over time, resulting in different structural characteristics (Fan et al., 2012).
The mixing of reducing groundwater and oxidizing surface water results in a sequence from oxygen reduction to nitrate reduction, ferromanganese reduction, sulphate reduction and then methanogenesis appears at the soil profile (Krause et al., 2011) (Fig. 2b). Therefore, SW-GW interactions can stimulate the functionally active chemolithotrophic bacteria (Storey et al., 1999), which extract energy from the oxidation of inorganic matter, such as sulphur, iron, nitrite and ammonia, and promote the oxidation-reduction reactions. When ammonia is absent, bacteria may use Mn2+, Fe2+ or reducing sulphur (elemental sulphur or sulphide) as an energy source. Denitrification converts nitrate to nitrogen over several steps, resulting in coupled oxidation of organic compounds in the absence of oxygen (Zhou N Q et al., 2014). With the exhaustion of oxygen, nitrate, iron, manganese and sulphate, and the accumulation of large amounts of CO2 (a by-product of most catabolic processes), methanogenesis occurs (Lee et al., 2009). Microbial biofilms play an essential role in this process. These biofilms are mainly composed of highly diverse bacterial and archaeal communities embedded in the same polysaccharide matrix (Battin et al., 2016).
Seasonal and daily fluctuations between low and high water levels cause variations in open channel discharge. Increased surface flow promotes more water and dissolved solutes input into the HZ, causing an increase in dissolved oxygen, while the water residence time in the HZ decreases. Also, biological effects should not be neglected. Biodisturbance and solute exchange occur in bottom sediments because of the activities of large invertebrates and small benthic fauna in areas of preferential flow and high permeability. The activities of large invertebrates and meiofauna lead to bioirrigation and bioturbation in streambed sediments, resulting in preferential flow paths and locally increased permeability. Protozoa grazing on biofilms increase their absorption surface. Therefore, the diffusion gradient of the dissolved solute in the presence of grazers is higher than when they are absent (Peralta-Maraver et al., 2018).
As an important part of the hydrological cycle, wetlands have prominent hydrological regulation functions. These not only slow down and restrain surface runoff, but also play a prominent role in unloading flood waters (Acreman and Holden, 2013). Wetland soils have special hydrophysical properties (e.g., grain size, hydraulic conductivity and organic matter content). These control the SW-GW interactions, the speed of water movement in the soil and the ability of the soil to absorb water. The presence of a macroporous preferential path allows water to move faster into the groundwater through the soil (Acreman and Holden, 2013). SW-GW interactions reduce flood peaks and make them discharge more smoothly and slowly, thus prolong the time that floods remain on land (Zedler and Kercher, 2005). Flood water can be released from wetlands in days, weeks or months; some increase local air humidity by evaporation during the flow process and some infiltrate to replenish groundwater and increase groundwater reserves. SW-GW interactions give wetlands the role of distributing and homogenizing river runoff.
Wetland vegetation can also slow down the flood flow, thus preventing all flood water from reaching downstream at the same time. In addition, based on root distribution through the soil profile, vegetation absorb water from different depths of the soil for evapotranspiration. Studies have shown that the utilization rate of groundwater by different plants is between 0 and 77% (O'Grady et al., 2006), and the absorption of groundwater by plants will cause a significant drop in groundwater level (Eamus et al., 2015), which will increase the surface water recharge to groundwater. Vegetation varies in its architecture and therefore in its efficiency as a conduit for water moving to the atmosphere from the soil.
Although there are many obvious examples of wetland flood reduction services, Bullock and Acreman (2003) showed that limited support for generalized models of wetland flood control. They demonstrated that about 80% of related researches showed that floodplain wetlands decreased flooding, but they also implied that 41% of headwater wetlands studies showed that these wetlands increased flooding. Acreman and Holden (2013) further indicated that upland wetlands generally tend to be flood generating areas while floodplain wetlands have a greater potential to reduce floods, and they also showed relative influence of wetlands with different management regimes on floods (Fig. 3). Figure 3 also showed the existence of permeable soils promoted the SW-GW interactions and therefore significantly reduce the flood magnitude.
The presence of saturated soils is one of the most important ecohydrological characteristics of wetlands. A water table near the soil surface creates an interaction of groundwater flow with a biologically active zone where plants and soil microbes dominate biogeochemical processes under anaerobic conditions (Xu et al., 2019; Millar et al., 2018), making this zone a mechanical filter composed of flowing water and porous medium and a biochemical filter governed by chemical and biological processes.
Vegetation can absorb elements from water or soil, such as nitrogen, phosphorus, potassium, calcium, sulfur, magnesium, iron, and others. This will indeed directly purify the pollution within the root system and change the geochemical gradients, and therefore indirectly result in chemical-biological processes in HZ (Moore, 2007). At the same time, vegetation plays critical role in carbon cycle in wetlands, such as sequestration, mineralization, storage, import/export and emissions (Luo M et al., 2017), which makes wetland serve as a biological pump in the carbon cycle and act as a bridge for microbial-oil-plant interactions in sediments.
Wetlands therefore play critical players in the flow of energy and biomass, nutrient cycles and pollution attenuation. Harvey et al. (2013) reported that denitrification began when surface water went underground and mainly happened in the shallow layer of the HZ. Clinton et al. (2010) concluded that the HZ can effectively degrade biological phosphorus in water through the adsorption of media and microorganisms. Methane production and loss caused by anaerobic metabolism in the HZ may be an important way to remove carbon from streams.
In recent years, research on micropollutants has received increasing attention, such as ibuprofen or antibiotics in water. Some studies have researched the degradation of micropollutants in wetlands and have shown that under certain conditions micropollutants can be effectively degraded along the flow path within the sediments (Fig. 4b) (Lewandowski et al., 2011). In fact, some of these chemical compounds (i.e., ibuprofen, diclofenac, naproxen and bezafibrate) have higher conversion efficiencies in sediments than by biofilms in wastewater treatment plants (Radke et al., 2009). This is mainly because of the higher diversity of microbial communities in the sediments environment. Moreover, the residence time of water in sediments is longer than in surface sediments and open channels, resulting in more efficient biodegradation (Lewandowski et al., 2011).
Wetland ecosystems contain high diversity and abundance of fauna. SW-GW interactions in wetlands result in a more stable living environment than groundwater and surface water in HZs. So a HZ is mainly home to invertebrates (mostly insect larvae and crustaceans), including stygobites (specialized to hypogean groundwater habitats), stygoxenes (occasionally underground) or stygophiles (epigean animals pre-adapted for underground life), and it is also a priority environment for spawning fish, such as salmon that lay their eggs in gravel (Bertrand et al., 2011). Changes in groundwater level and surface water flow caused by alternating wet and dry seasons may change the water exchange in HZs and affect this biome. However, the HZ maintains humidity after surface drying and remains stable during flooding, so these potential effects can be reduced in the HZ (Robertson and Wood, 2010). When groundwater and surface water change, the HZ may serve as a refuge for local biota and, and when the environment stabilizes, these organisms can return to their original habitats (Dole-Olivier, 2011). In summer, the HZ temperature is lower than the surface environment, so it can serve as a refuge for some organisms whose biological processes, such as membrane transport, diffusion and enzymatic reactions, are governed by the temperature of the local environment (Zhou N Q et al., 2014). The HZ can therefore enhance the resilience of benthic communities and affect the restoration of wetlands. In addition, pore size and oxygen concentration reduce with depth because of HZ processes (Fig. 2). As a result, with increasing depth the invertebrate assemblage, comprising a relatively small group of large individuals near the surface, are replaced by numerous smaller organisms. The density of large macroinvertebrates declines, while that of meiofauna and protists increases (Fig. 5a). So that the scope and structure of a HZ in wetlands have significant influence on the geographic distribution and composition of benthic communities (Krause et al., 2011).
Biogeochemical processes within a HZ are vital to maintain the cycling of carbon and nutrients, and maintain the stability of the wetland food chain (Feris et al., 2003). Dissolution and microbial transformation of particulate nutrients in HZ have an impact on macroinvertebrate and algal assemblages, and may influence the productivity of riparian vegetation (Clarke, 2002). Within the HZ, biogeochemical cycles, microbial and animal ecology should be studied as an integrated unit. A typical flow of a carbon-based HZ food web composed of the microbial loop and grazer chain is illustrated in Fig. 5b. As tiny (50-1 000 µm) interstitial invertebrates, meiofauna act as a carbon conduit from large invertebrate consumers to microbial biofilms (Hancock et al., 2005). The connection of this prey-predator between meiofauna and bacteria has been detected in a HZ. When large invertebrates directly consume particulate organic matter, they enhance bacterial activity by providing nutrients through death or excretion, increasing the surface area available for attack and raising the flux of oxygenated water by tunneling (Hancock et al., 2005).
The heterogeneity and spatial-temporal variability of SW-GW interactions make the study of wetlands complicated, and many methods have been applied to them. Kalbus et al. (2006) provided a detailed review of these conventional methods. This paper summarizes their reviews, and focuses on the investigation methods used in geophysics that have emerged in recent years.
A few techniques allow for mapping SW-GW interactions (Table 1). The most direct approach is to calculate water exchange intensity and mixing ratio, and to divide the scope of HZs by measuring head pressure and chemical components at different sample points and depths. However, this method has a large workload and is limited to small-scale wetlands (Battin et al., 2009). Similarly, the tracer-injection method can directly obtain the depth of the HZ and the residence time by analysing the breakthrough curve of the tracer. Although this method is more effective, it takes a long time and is also controlled by the selection and arrangement of monitoring points (Du et al., 2017a). There are also different methods of using environmental tracers, based on the content of non-reactive environmental tracers in different water bodies (surface water, groundwater, precipitation, etc.) and the concentration gradient of the radiotracer in the sediment pore water (Luo X et al., 2017). Thermal tracing has become a popular method because of the ubiquitous changes of heat (including day and night changes and seasonal changes), the developments in temperature measurement equipment and the progress in heat simulation methods. But, at a large scale, the heat signal is often blurred and the temperature monitoring rarely exceeds the decimetre scale (House et al., 2015).
| Methods | Different types | Principle |
| Direct measurements | Water flux measurement | Applying bag-type seepage meters to measure the water flow |
| Head pressure measurement | Based on calculation of gradients and conductivity and Darcy's Law | |
| Tracer | Hydrochemical indicators | The hydrochemical characteristics of different water are obviously different under the influence of different sources and different migration paths |
| Isotope tracers | Similar as the hydrochemical indicators | |
| Tracer test | Injecting conservative compounds or isotopes to identify SW-GW interactions | |
| Thermal tracing | Point temperature measurement | Using traditional glass thermometers, pressure sensors, thermistors, thermocouple temperature probes, optical and radiation sensors, etc. to get continuous or discrete temperature data at different points |
| Distributed temperature sensing | Temperature values at different spatial points are measured according to the distribution of light at different wavelengths and the time when the reflected signal is received | |
| Thermal infrared remote sensing temperature measurement | Using satellite or airborne sensors to collect and record the thermal infrared information of terrestrial objects, and then to retrieve the surface temperature |
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In recent years, as a new investigation, geophysical exploration methods have been applied to the identification of SW-GW interactions. The most commonly used method is electrical resistivity imaging (ERI). This method is based on the difference of salt content of groundwater and surface water which will lead to different resistivity, and the interaction relationship between them can be analysed by resistivity map. This method can obtain 2D or 3D resistivity distribution maps at different times (Befus et al., 2012), and then infer the structural characteristics and scope of the HZ, including the water exchange characteristics (Busato et al., 2019; Cardenas and Markowski, 2011). However, the ERI method often requires a significant difference in resistivity between surface water and groundwater, which makes this method more suitable for the study of coastal wetlands (Day-Lewis et al., 2006). However, some studies have researched its application in surface water-groundwater exchange in non-coastal areas by repeated measurements at different time and adding tracers (Nyquist et al., 2008; Singha et al., 2008).
The ERI method can be used to detect different time points and accuracy requirements, allowing both chemical and physical mixing processes to be unraveled in ubiquitous environments whose study has been challenging. It can also be used to evaluate 3D and larger scale patterns, making it possible for HZ scope and structure to be mapped across nested scales (Cardenas and Markowski, 2011), which gives this method wide potential application. The spatial resolution of ERI-derived tomographic images is affected by survey geometry, regularization, prior information, measurement sensitivity and errors and inversion constraints (Day-Lewis et al., 2006).
Many researchers have reviewed the methods of mathematical simulation of SW-GW interactions, such as Boano et al. (2014), Barthel and Banzhaf (2015) and Brunner et al. (2017). Overall, according to these differences, the existing models can be classified into two categories. One is a physical model based on physics, and the other is a model based on phenomenology (Boano et al., 2014). The physical model is based on the principles of mass and momentum balance to determine the driving force and volume of surface water and groundwater exchange. It can be used to describe the control factors and exchange dynamics of water flow in the subsurface flow zone. However, such models often require large amounts of data and parameters, such as permeability coefficient and surface water-groundwater dynamics, which limit the ability of the model to predict water flow. Based on the phenomenological descriptions of SW-GW interactions, a large number of models have also been constructed, the most famous of which is the transient storage model (Liao et al., 2013). Developed from the classical advection-dispersion model, the transient storage model mainly explains complex physical processes occurring at a small scale.
More mature and wide applicability models applied to evaluate the SW-GW interactions regionally can further classified into two types (Barthel and Banzhaf, 2015): fully coupled schemes whose equations governing surface water and groundwater flows are solved simultaneously within one software package (e.g., ParFlow, HydroGeoSphere, InHM, and OpenGeoSys) and loosely coupled schemes which are coupled by two or more individual models via the exchange of model results, where the output of one model forms the input of the other (e.g., DAFLOW, MD-SWAT-MODFLOW, DYNSYSTEM, and IGSM). Many of these models can not only calculate the water exchange between surface water and groundwater, but also to evaluate the solutes transport and heat exchange (Maxwell and Condon, 2016; Maxwell et al., 2015).
With changes in precipitation patterns, the greatest increase in runoff occurs during wet seasons, whereas the greatest decrease is possible during dry seasons (Zhou S B et al., 2014). In wet seasons, intensified runoff will carry increased sediments and pollution, which would increase clogging in HZs. However, in dry seasons, the combined effects of reduced precipitation and high evaporation caused by rising temperatures will lead to a lower water level. The decrease of water resource has caused shrinkage of wetland area, salinization and wetland succession in many wetlands globally (Nachshon et al., 2014; Nielsen and Brock, 2009). The annual distribution of precipitation, freezing and snow melt, rainy season flooding and strong evapotranspiration in windy seasons lead to significant differences in SW-GW interaction dynamics in wet and dry seasons (Sophocleous, 2002).
Higher temperatures in winter will reduce ground frost and make more water penetrate into the ground, thus increasing groundwater recharge (Kløve et al., 2014). Increased aquifer recharge will raise the winter groundwater level, while in spring and summer, the groundwater level may be lower than average water tables (Okkonen and Kløve, 2010). A warmer winter would also cause earlier snow melting, which further shifts flood peak earlier and reduces summer flow (Okkonen and Kløve, 2011). Higher temperatures will increase convection and evaporation, indirectly resulting in reduced water level in some regions (Leigh et al., 2013). Ficklin et al. (2013) projected that warmer temperature could increase average annual evapotranspiration by nearly 23% by the end of the twenty-first century in the Colorado River Basin of the USA. Changes in surface water and groundwater levels may eventually change SW-GW interactions and the interaction between natural and social water demand and supply (Hanson et al., 2012). As the temperature increases, the spatial expansion of wetlands may decrease as the groundwater and surface water levels decrease.
As the amount of precipitation decreases, summer droughts become more frequent and dense, which reduces the water level of the wetland. The lower water level can introduce oxygen into wetland deposits, accelerate decomposition and destroy the previously stable wetland environment (Moore et al., 2013). Melting glaciers and more frequent storms wash away the soil, transporting large amounts of particulate and dissolved organic material, stimulating microbial metabolism and HZ respiratory activities (Thangarajan et al., 2013).
Temperature dynamics directly affect organism activity and heterogeneity in a HZ (Dole-Olivier, 2011), and indirectly change O2 availability and organic matter supply (Wang R C et al., 2018; Peyrard et al., 2011). An increase in average temperature will speed up water exchange and decomposition of organic carbon in sediments, thereby delivering greater amounts of CO2, CH4 and/or DOC depending on the edaphic conditions (Mitsch et al., 2012). Particularly, this effect may be most pronounced in swamps or peatlands in Arctic regions, as the decomposition of these zones is usually limited by low temperatures, which are expected to rise the most in high latitudes of the northern hemisphere. Both long-term observations and experimental research have suggested that carbon pools in tundra or peatland ecosystems may become new carbon sources to the atmosphere due to warmer temperatures (Yu, 2012).
During dry seasons, lower water exchange will result in interstitial clogging, permeability decrease and substance exchange in the sediments. The sediment permeability has a greater effect on bacterial diversity and composition than temperature (Zeglin et al., 2011). There is a positive correlation between hyporheic species survival and dissolved organic material provided by SW-GW interactions, but the low permeability will limit the food sources (Buendia et al., 2013). The HZ supplies an ideal area for fish embryos and protects them from predators, but clogging of a streambed has been reported to damage reproductive success and rates of fish recruitment (Scheurer et al., 2009). So that the loss of susceptible taxa resulting from the prolongation of the dry season can significantly alter assemblages of hyporheic invertebrates (Datry, 2012).
During wet seasons, increased runoff from surrounding residential area carries high concentrations of pathogens and contaminants into wetland water bodies, favoring the growth and reproduction of pathogens indirectly. In addition, higher temperatures enhance the emergence and persistence of many infectious microorganisms (Harvell et al., 2009). The reproduction of pathogens may affect the structure of the original biofilms in the sediments and the environmental tolerance of individual species. An explosive bacterial growth combined with reduced biofilm resistance may affect the food supply of the macroinvertebrates and further the wetland health.
Major anthropogenic causes of wetland degradation include waste disposal, agricultural and forestry drainage, flood control and stream channelization, development of roads or urban settlements, water diversion irrigation, groundwater extraction, fertilizer and sewage runoff and pesticide leakage, exploitation of minerals, peat or gravel and sea level rise (Moore, 2007). Such factors may directly or indirectly influence sedimentation and hydrological regimes, water quality and biological resources.
Urbanization generally increases total flow and peak runoff by converting wetland soils into impervious surfaces and impairing water quality. Because of an increased water demand in urbanized areas, dam construction and water abstraction will also affect SW-GW interactions in wetlands by modifying the frequency and level of environmental flow. Reduction of hydrological flow typically leads to sedimentation instream and declining depth to the extent that vegetation (especially exotics) invade shallower parts (Lee et al., 2006). Expansion of multi-lane freeways and highways has led to restrictions on river migration, the dumping of waste and toxic substances into rivers and reduction of wetland and riparian habitats adjacent to the roadway (Patten, 2006).
Historically, wetlands have been drained for the development of farmland globally. Such drainage is still ongoing and even increasing in scale, especially in areas where wetland drainage is often utilized as a way of increasing biomass or crop production. The direct use of wetland surface water will lead to reductions in water storage and the shrinkage of wetlands. The formation of a depression cone from groundwater overexploitation for irrigation will intensify the drainage of wetlands to groundwater (Johansen et al., 2011). The water channels can directly supply water for wetlands, playing a crucial role in the process of wetland restoration, and then promoting the rise of the groundwater level, water quality and vegetation growth (Richardson et al., 2011). However, by constructing channels, ditches and canals, or straightening streams and removing natural barriers, such as vegetation, wetland outflow will also increase (Rasmussen et al., 2018), and surface water would recharge more to groundwater (Wu et al., 2017). Recent research has demonstrated that in tropical wetlands, the construction of ditches to reduce water content causes DOC release from paleo-organic carbon stocks (Moore et al., 2013).
The migration and enrichment of large amounts of nitrogen, phosphorus, heavy metals and organic matter from anthropogenic pollutions leads to deterioration in wetland water quality (Ma et al., 2016). The biogeochemical behavior of nitrogen, phosphorus and sulphur derived from wetlands has a negative impact on the evolution of groundwater chemistry. In the context of human activities leading to widespread deterioration of the environment, the impact of polluted groundwater on wetland systems is also a current research hotspot.
As a way of wetland restoration, constructed wetlands are well developed all over the world. There are two regimes to develop a constructed wetland: optimizing the existing wetlands or creating wetlands where there were none previously (Mateos, 2017). These wetlands are often constructed to achieve multi-functionality rather than focusing on specific ecosystem services. These services include biodiversity enhancement, soil desalinization and water quality improvement (Moreno-Mateos and Comín, 2010). Today, most constructed wetlands are designed to eliminate organic compounds, nutrients and heavy metals from both urban and agricultural wastewater as efficiently as possible (Moreno et al., 2007). Under highly controlled conditions, constructed wetlands can achieve high long-term removal rates of organic compounds, phosphorus and nitrogen by manipulating plant species, flow speed, substrates, oxygenation and microbial communities. Many literatures has demonstrated that wetlands were constructed to treat contaminated groundwater, and the results showed that nitrate and organic matters can be effectively removed from groundwater because of the SW-GW interactions (Coban et al., 2015; Seeger et al., 2011).
Field studies usually suffer from an absence of chemical and ecological baseline data to describe, evaluate and quantify SW-GW interactions processes and functions over a range of scales. It is necessary to develop new, alternative, more efficient and robust methods and techniques to obtain long-term sequence data on hydrology, chemistry and biology for future research. Detailed studies of hyporheic processes are often limited to small-scale research, and monitoring and research on larger scales is greatly simplified. So that it is also necessary to conduct monitoring at different spatial and time scales, and to raise the accuracy and resolution of the monitoring. Also, these simulation methods are partially coupled with water exchange, physical adsorption and deposition, chemical processes and biological metabolism, but there is no model that fully couples all processes. In addition, the existing models are not applicable for the complex multi-scale characteristics of hyporheic exchange, nor can they solve the problem of upscaling or downscaling the prediction results.
The continuum of SW-GW interactions in wetlands is broad, with spatial scales ranging from millimeters or centimeters, beneath small bed forms, to longer flow paths of tens, hundreds, or even thousands of meters and temporal scales ranging from seconds to tens of years. It is affected by many factors, such as hydrological conditions, hydrogeological conditions, geomorphological conditions, meteorological and climatic conditions, which involve surface water velocity, discharge, sediment transport, sediment permeability, saturation rate, surface water and groundwater head difference, river bed slope, bay shape, rainfall and evaporation, temperature change and so on. These variables have strong variability and various influences in different spatial and temporal scales, making it difficult to study these variables (Boano et al., 2014). In addition, to apply the results of more detailed small-scale studies to large-scale studies, new techniques and methods need to be developed to determine the controlling effects of spatial and temporal variables on HZ processes. How to assess the influence of these factors on wetland-groundwater interaction under different conditions, find out the key variables, and further study the threshold value of these variables and combinations of multiple variables will be an important topic in future research.
There is an urgent need to analyze the dynamic changes in wetland water cycles, distribution of water resources and the evolution of the water environment based on future climate change and human activity scenarios, and to identify the response mechanism of SW-GW interactions in wetlands under a changing environment. An integrated water resources management scheme based on this needs to be proposed to maintain the health and stability of WEs (Kløve et al., 2014). The scheme demands coordinated development of land, water and related resources to maximize the resultant social and economic welfare in an equitable manner, without compromising the sustainability of vital ecosystems. To be fully effective, groundwater management within the general integrated water resources management framework requires integration of institutional arrangements, an appropriate regulatory and policy framework, economic instruments and social participation.
As an essential type of ecosystems, various types of wetlands depend on groundwater for stability and sustainability. Groundwater provides a reliable source of water, solutes, nutrients and energy for these wetlands and has become the focus of research on wetland ecohydrology. Wetland HZs are one of the most important transition zone of surface water and groundwater, and are also the key to studying mechanisms and processes of SW-GW interactions. The differences of groundwater and surface water result in significant physical, chemical and biological gradient in HZ, and promote material reactions and energy cycling between the groundwater and surface water. Therefore, SW-GW interactions are crucial for wetlands to maintain their ecological functions, such as regulation of water resources, removing pollutions and maintenance of biological diversity. However, climates changes and human activities have significantly altered the SW-GW interactions and degraded the function of wetlands.
Although the functions of groundwater in wetlands are important, previous research into groundwater hydrological processes in wetlands has been inadequate and there has not been intensive research into the interaction mechanisms between surface water and groundwater and their ecological effects. Hence, questions on how wetlands depend on the groundwater system to maintain their stability, how global climate change and human activity affect SW-GW interactions in wetlands and how humans should regulate SW-GW interactions to ensure the sustainable development of wetland ecosystems remain unanswered.
Therefore, the studies of SW-GW interactions in wetlands should be intensified. Integrated multidisciplinary monitoring and simulation methods should be developed to get the chemical and ecological baseline data and to simulate physical-chemical-biochemical and ecological processes, and further to study the dominant variables and thresholds of SW-GW interactions to assess the importance of groundwater in wetlands. Finally, researchers need to conduct integrated water resources management program to realize the harmonious development of wetland ecology and social economy under the influence of climate change and human activity.
This project was supported by the National Natural Science Foundation of China (Nos. 41630318, 41521001), the Project of China Geological Survey (Nos. 121201001000150121, DD20190263, 2019040022), the Research Program for Geological Processes, Resources and Environment in the Yangtze River Basin (No. CUGCJ 1702). We thank Kara Bogus, PhD, from Liwen Bianji, Edanz Editing China (www.liwenbianji.cn/ac), for editing the English text of a draft of this manuscript. The final publication is available at Springer via https://doi.org/10.1007/s12583-020-1333-7.
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Figures(5) / Tables(1)
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Xiancang Wu, Teng Ma, Yanxin Wang. Surface Water and Groundwater Interactions in Wetlands. Journal of Earth Science, 2020, 31(5): 1016-1028. doi: 10.1007/s12583-020-1333-7
| Methods | Different types | Principle |
| Direct measurements | Water flux measurement | Applying bag-type seepage meters to measure the water flow |
| Head pressure measurement | Based on calculation of gradients and conductivity and Darcy's Law | |
| Tracer | Hydrochemical indicators | The hydrochemical characteristics of different water are obviously different under the influence of different sources and different migration paths |
| Isotope tracers | Similar as the hydrochemical indicators | |
| Tracer test | Injecting conservative compounds or isotopes to identify SW-GW interactions | |
| Thermal tracing | Point temperature measurement | Using traditional glass thermometers, pressure sensors, thermistors, thermocouple temperature probes, optical and radiation sensors, etc. to get continuous or discrete temperature data at different points |
| Distributed temperature sensing | Temperature values at different spatial points are measured according to the distribution of light at different wavelengths and the time when the reflected signal is received | |
| Thermal infrared remote sensing temperature measurement | Using satellite or airborne sensors to collect and record the thermal infrared information of terrestrial objects, and then to retrieve the surface temperature |
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