Groundwater use and risks in the coastal zone

As surface water resources become more limited and progressively degraded, coastal groundwater abstraction has increased markedly and is expected to increase even further under predicted drier conditions in the future with more frequent, long-lasting and intense droughts anticipated (e.g., Parisi et al., Reference Parisi, Monno and Fidelibus2018; Erostate et al., Reference Erostate, Huneau, Garel, Ghiotti, Vystavna, Garrido and Pasqualini2020). These climatic changes (higher temperatures, evapotranspiration and decreased precipitation) will further affect groundwater resources through decreased recharge (e.g., Stigter et al., Reference Stigter, Nunes, Pisani, Fakir, Hugman, Li, Tomé, Ribeiro, Samper, Oliveira, Monteiro, Silva, Tavares, Shapouri, Cancela da Fonseca and El Himer2014; Unsal et al., Reference Unsal, Yagbasan and Yazicigil2014). The consequent impacts of declines in precipitation and groundwater recharge are expected to be especially pronounced in arid and semi-arid regions and coastal regions, such as small islands, that rely exclusively on rainwater harvesting (RWH) and groundwater resources for water supply (e.g., Stigter et al., Reference Stigter, Nunes, Pisani, Fakir, Hugman, Li, Tomé, Ribeiro, Samper, Oliveira, Monteiro, Silva, Tavares, Shapouri, Cancela da Fonseca and El Himer2014; Unsal et al., Reference Unsal, Yagbasan and Yazicigil2014; Kopsiaftis et al., Reference Kopsiaftis, Tigkas, Christelis and Vangelis2017). Urbanisation may further decrease groundwater recharge by the replacement of natural permeable landscapes with hard, impermeable surfaces such as roads. These urban surfaces linearly direct surface runoff to lentic waterways more effectively and therefore reduce infiltration opportunities (Han et al., Reference Han, Currell, Cao and Hall2017; Lorenzo-Lacruz et al., Reference Lorenzo-Lacruz, Garcia and Morán-Tejeda2017; Minnig et al., Reference Minnig, Moeck, Radny and Schirmer2018; Frommen et al., Reference Frommen, Groeschke, Nölscher, Koeniger and Schneider2021). As such, this will contribute to the coastal squeeze of coastal urban aquifers since these systems are wedged between developed urban areas and rising seas (Vitousek et al., Reference Vitousek, Barnard and Limber2017).

Coastal threats and changes (e.g., SLR, storm surges, coastal erosion and land subsidence) may occur simultaneously and/or have knock-on effects leading to the occurrence or acceleration of other pressures, usually intensified through urbanisation and development. The onset of coastal pressures may, however, also be triggered inland such as in the case of coastal erosion caused by dam construction and the resultant changes in sediment source and coastal sediment dynamics (e.g., Huang and Jin, Reference Huang and Jin2018 and references therein; Hzami et al., Reference Hzami, Heggy, Amrouni, Mahé, Maanan and Abdeljaouad2021). Coastal erosion may accelerate land losses beyond those caused by SLR, or erosion may amplify the effects of SLR, causing groundwater salinisation and reductions in groundwater recharge (Baharuddin et al., Reference Baharuddin, Masirin, Hazreek, Azman and Madun2018; Chesnaux et al., Reference Chesnaux, Marion, Boumaiza, Richard and Walter2021). This is of further concern since the decrease in groundwater recharge may be a more important issue than SLR in flat low-lying coastal areas (Parisi et al., Reference Parisi, Monno and Fidelibus2018). Diminished groundwater discharge levels due to declines in groundwater levels, whether induced by abstraction or reductions in recharge, affect the rest of the hydrologic system through decreased river baseflows, lake and spring levels and ultimately result in landscape desiccation (de Graaf et al., Reference De Graaf, Gleeson, (Rens) Van Beek, Sutanudjaja and Bierkens2019). Furthermore, seaward erosion during storm surges of coastal barriers (e.g., dunes), could result in breaching events that may subsequently affect coastal aquifers kilometres inland through vertical saltwater infiltration, with the delayed recovery of the aquifers to a freshwater state lasting longer than the storm event (Giambastiani et al., Reference Giambastiani, Colombani, Greggio, Antonellini and Mastrocicco2017; Elsayed et al., Reference Elsayed, Oumeraci and Goseberg2018).

Many coastal hazards, whether caused by anthropogenic or climatic drivers, ultimately lead to coastal inundation (permanent) or flooding (temporary) (sensu Flick et al., Reference Flick, Chadwick and Harper2012) (hereafter collectively referred to as submergence). Three types may be recognised, namely: 1) marine submergence (surface submergence with marine water through, e.g., tidal flooding, SLR, storm drain backflow); 2) groundwater submergence (rise in groundwater levels as groundwater is displaced in response to, e.g., SLR or wastewater network leaks) and 3) precipitation-driven submergence (through large rainfall events) (Han et al., Reference Han, Currell, Cao and Hall2017; Minnig et al., Reference Minnig, Moeck, Radny and Schirmer2018; Habel et al., Reference Habel, Fletcher, Anderson and Thompson2020; Rahimi et al., Reference Rahimi, Tavakol-Davani, Graves, Gomez and Valipour2020; Su et al., Reference Su, Liu, Beheshti and Prigiobbe2020; Frommen et al., Reference Frommen, Groeschke, Nölscher, Koeniger and Schneider2021). Furthermore, these submergence types may compound each other, which can result in “snowballing” impacts compared to isolated occurrences (Habel et al., Reference Habel, Fletcher, Anderson and Thompson2020; Rahimi et al., Reference Rahimi, Tavakol-Davani, Graves, Gomez and Valipour2020). For example, modelled scenarios of the Oakland Flatlands, a low-lying coastal area of California, USA indicate that SLR alone will not pose significant threats to urban infrastructure, but coupled with groundwater submergence more than 310 ha will be negatively affected (Rahimi et al., Reference Rahimi, Tavakol-Davani, Graves, Gomez and Valipour2020). Excluding the damage to coastal infrastructure and threats to human life, submergence is usually accompanied by groundwater salinisation through seawater intrusion and pollution of aquifers (Habel et al., Reference Habel, Fletcher, Anderson and Thompson2020; Rakib et al., Reference Rakib, Sasaki, Matsuda, Quraishi, Mahmud, Bodrud-Doza, Atique Ullah, Fatema, Newaz and Bhuiyan2020), which exacerbates the degradation of aquifers already under pressure from over-abstraction and droughts by decreasing potability (e.g., in California, USA – Rahimi et al., Reference Rahimi, Tavakol-Davani, Graves, Gomez and Valipour2020).

Despite being identified as a hazard to drinking water supply systems more than a century ago (Michael et al., Reference Michael, Post, Wilson and Werner2017), seawater intrusion and resultant groundwater salinisation are still the primary threat related to coastal groundwater abstraction (Chang et al., Reference Chang, Zheng, Zheng, Zhang, Sun and Walther2019). The source of salinity, however, may be natural (seawater trapped in pores from earlier sea level transgressions and dissolution of evaporite minerals, e.g., halite and gypsum) or anthropogenic (related to over-extraction and agricultural return flow) (Nogueira et al., Reference Nogueira, Stigter, Zhou, Mussa and Juizo2019).

Seawater intrusion triggered by over-abstraction from coastal aquifers can result in both landward (lateral) and vertical (upconing) intrusion (Parisi et al., Reference Parisi, Monno and Fidelibus2018). Coastal aquifers are increasingly at risk from bidirectional contamination as shorelines migrate landward with SLR and seawater intrusion and with pollution pressure from land sources – especially agricultural and industrial (Michael et al., Reference Michael, Post, Wilson and Werner2017; summarised in Figures 1 and 2). Not only does this leave groundwater resource unsuitable for human use but may also have dire effects on coastal groundwater-dependent ecosystems (GDEs) such as estuaries, lagoons, microbialite ecosystems and coral reefs (Michael et al., Reference Michael, Post, Wilson and Werner2017; Erostate et al., Reference Erostate, Huneau, Garel, Ghiotti, Vystavna, Garrido and Pasqualini2020; Rishworth et al., Reference Rishworth, Dodd, Perissinotto, Bornman, Adams, Anderson, Cawthra, Dorrington, Du Toit, Edworthy, Gibb, Human, Isemonger, Lemley, Miranda, Peer, Raw, Smith, Steyn, Strydom, Teske and Welman2020b). These GDEs are also at risk of state changes or even localised extinction as groundwater levels are reduced or depleted and surface-groundwater connections are disturbed. This is especially relevant in arid or semi-arid coastal regions with a Mediterranean climate, such as those of the Mediterranean basin and the southwestern coasts of Australia, Chile, California (USA) and South Africa (Erostate et al., Reference Erostate, Huneau, Garel, Ghiotti, Vystavna, Garrido and Pasqualini2020). For example, salinity increases in the microbialite-bearing Lake Clifton, Western Australia has been observed since the 1990s. This is attributed to hydrological changes driven by increased evaporation and decreased freshwater recharge to Lake Clifton due to lower rainfall, groundwater abstraction in the catchment areas and the construction of the Dawesville Channel (Forbes and Vogwill, Reference Forbes and Vogwill2016; Warden et al., Reference Warden, Coshell, Rosen, Breecker, Ruthrof and Omelon2019). The higher salinity, in addition to higher nutrient loads, is thought to have contributed to a change in the microbial communities of the Lake, although this does not necessarily translate to unfavourable conditions for microbialite formation (Gleeson et al., Reference Gleeson, Wacey, Waite, O’Donnell and Kilburn2016; Warden et al., Reference Warden, Coshell, Rosen, Breecker, Ruthrof and Omelon2019). The effects of coastal water crises associated with groundwater salinisation globally may be summarised as conflict of uses between different water users, wetland degradation, loss of crops, soil salinisation, aridity, desertification and health issues (Parisi et al., Reference Parisi, Monno and Fidelibus2018).

Figure 2. Driver-Pressure-State-Impact-Response (DPSIR) framework regarding groundwater use in coastal urban areas. Responses with a cross indicate there are large negative implications reported with the use thereof.

Furthermore, an emerging issue related to coastal groundwater use is that of land subsidence and increased rates of relative SLR. Natural land subsidence of unconsolidated coastal sediments may be accelerated by groundwater abstraction and urban development (see below) and when combined with SLR and storm surges increase the risk of seawater intrusion into aquifers (Huang and Jin, Reference Huang and Jin2018; Shirzaei et al., Reference Shirzaei, Freymueller, Törnqvist, Galloway, Dura and Minderhoud2021). This will likely be exacerbated by drought conditions and climate change. For example, model predictions underestimated land subsidence in Yuanchang, Taiwan for the 2011–2012 period due to increased groundwater pumping and reduced recharge during a dry spell (Shirzaei et al., Reference Shirzaei, Freymueller, Törnqvist, Galloway, Dura and Minderhoud2021). Relative SLR may pose a further risk to coastal communities through groundwater flooding as unconfined coastal aquifer levels will increase in response to sea level changes. It is anticipated that water tables will either rise the same amount as sea levels, where the additional groundwater can be accommodated, or that the seawater intrusion will displace fresh groundwater, which will discharge through original or new drainage systems (Befus et al., Reference Befus, Barnard, Hoover, Hart and Voss2020). For instance, coastal cities in Indonesia such as Semarang and the Indonesian capital, Jakarta, are already experiencing the impacts of land subsidence, coastal flooding during high tides and seawater intrusion into aquifers (Abidin et al., Reference Abidin, Andreas, Gumilar, Sidiq and Gamal2015). This is partly due to the demand for clean water outstripping surface water supply and consequently the over-abstraction of groundwater to make up the water supply deficit. Furthermore, urban development results in the consolidation of sediments, increased surface runoff and decreased groundwater recharge during rain events (Abidin et al., Reference Abidin, Andreas, Gumilar, Sidiq and Gamal2015; Pramono, Reference Pramono2021; Taftazani et al., Reference Taftazani, Kazama and Takizawa2022). As a result, more than 80% of groundwater samples from monitoring wells in northern Jakarta’s groundwater basins exceed the recommended limits of salinity for human consumption (Taftazani et al., Reference Taftazani, Kazama and Takizawa2022) and aquifers of entire small islands are brackish or saline during dry seasons (Cahyadi, Reference Cahyadi2018).

Water quality degradation through anthropogenic and geogenic pollution linked to increased water demand will further decrease both surface and groundwater resources available for use at the coast (e.g., Han and Currell, Reference Han and Currell2022). Urban aquifers are often associated with pollution by emerging contaminants (e.g., pharmaceuticals and personal care products) and nutrient loading through aqua- and agricultural activities, septic tanks, cemeteries and landfills (e.g., Lapworth et al., Reference Lapworth, Nkhuwa, Okotto-Okotto, Pedley, Stuart, Tijani and Wright2017; Boretti and Rosa, Reference Boretti and Rosa2019; Burri et al., Reference Burri, Weatherl, Moeck and Schirmer2019; Preziosi et al., Reference Preziosi, Frollini, Zoppini, Ghergo, Melita, Parrone, Rossi and Amalfitano2019; Han and Currell, Reference Han and Currell2022). Slow-moving aquifers and groundwater-fed rivers can release persistent organic pollutants, such as anthropogenic per- and polyfluoroalkyl substances (PFAS), into marine environments long after its initial release into the upstream environment (Sunderland et al., Reference Sunderland, Hu, Dassuncao, Tokranov, Wagner and Allen2019; Zhang et al., Reference Zhang, Lohmann and Sunderland2019a; Ruyle et al., Reference Ruyle, Pickard, Leblanc, Tokranov, Thackray, Hu, Vecitis and Sunderland2021). Coastal point sources include airports and ports (Li et al., Reference Li, Zhang, Gibbes, Wang and Lockington2022), while wastewater treatment works, landfills and urban runoff also contribute to the transportation of PFAS (e.g., Hepburn et al., Reference Hepburn, Madden, Szabo, Coggan, Clarke and Currell2019; Cui et al., Reference Cui, Li and Quinete2020). The residence time of certain PFAS species may be further lengthened by adsorption driven by tidal and salinity effects in the coastal zone resulting in bioaccumulation in benthic organisms. In fact, seafood is considered the primary human dietary exposure pathway to PFAS compounds (Sunderland et al., Reference Sunderland, Hu, Dassuncao, Tokranov, Wagner and Allen2019; Zhang et al., Reference Zhang, Lohmann and Sunderland2019a). However, PFAS adsorption in coastal sandy aquifers may act as a natural attenuation process (Li et al., Reference Li, Zhang, Gibbes, Wang and Lockington2022). These pathways are therefore not fully understood but present an emerging research priority and likely threat to coastal aquifers.

Management of water crises and groundwater use in the coastal zone

When facing a water scarcity crisis, there are two management strategies that can be employed, namely, supply management (i.e., increase the available resources) and demand management (i.e., decrease water use) (Lam et al., Reference Lam, Lant, O’Brien and Kenway2016). Although droughts occur naturally as a part of the hydrological cycle, their frequency and intensity are increasing under climate change scenarios. Emergency procedures during droughts often mean that more wells are drilled, which may accelerate aquifer depletion (Petersen-Perlman et al., Reference Petersen-Perlman, Aguilar-Barajas and Megdal2022) and are therefore likely not a sustainable solution (Figure 2). Water supply management in urbanised coastal areas during droughts cannot rely on a “wait ‘til it rains’” approach (Parisi et al., Reference Parisi, Monno and Fidelibus2018). Rather, socio-ecological and collaborative strategic management of water supply is required to effectively manage water resources and prevent aquifer salinisation (Parisi et al., Reference Parisi, Monno and Fidelibus2018).

Some problems related to drought and groundwater management are caused by governance misalignments (Petersen-Perlman et al., Reference Petersen-Perlman, Aguilar-Barajas and Megdal2022). For example, the Salento aquifer in Italy is experiencing groundwater salinisation and despite droughts in the area, salinity effects are not considered in groundwater management practices and no systematic monitoring takes place (Parisi et al., Reference Parisi, Monno and Fidelibus2018). Furthermore, although the national legislation of selected water-scarce and drought-prone countries protects GDEs (e.g., Australia) or advocates for environmental requirements (e.g., South Africa) (Rohde et al., Reference Rohde, Froend and Howard2017; Erostate et al., Reference Erostate, Huneau, Garel, Ghiotti, Vystavna, Garrido and Pasqualini2020), monitoring of actual regional groundwater abstraction is often lacking or management regulations are not met (Bekesi et al., Reference Bekesi, McGuire and Moiler2009; Erostate et al., Reference Erostate, Huneau, Garel, Ghiotti, Vystavna, Garrido and Pasqualini2020; Kent et al., Reference Kent, Pandey, Turner, Dickinson and Jamieson2020; Robertson, Reference Robertson2020). That being said, in Australia progress in GDE management has been achieved through comprehensive adaptive management frameworks that are informed by legislation as well as scientific and technological advancements coupled with regional datasets (Rohde et al., Reference Rohde, Froend and Howard2017).

Other accounts of successful groundwater management strategies and implementation of policy during droughts in coastal urban areas include the employment of managed aquifer recharge (MAR). The main applications of MAR are to increase water storage and provide a water resource more resilient to climate change, while water quality is also often improved by pollutant attenuation during infiltration (e.g., Kazakis, Reference Kazakis2018; Dillon et al., Reference Dillon, Escalante, Megdal and Massmann2020). This strategy of MAR is often implemented in regions with seasonal precipitation cycles to mitigate surface-water shortages during the dry months, such as in India (Glendenning and Vervoort, Reference Glendenning and Vervoort2011). A small-scale example, stemming from ancestral knowledge, is that of coastal Ecuador using constructed “tapes” or artisanal dikes to increase groundwater storage during periods of water scarcity (Carrión et al., Reference Carrión, Herrera, Briones, Sánchez and Limón2018). Furthermore, MAR has been widely applied to prevent seawater intrusion into coastal aquifers by maintaining groundwater levels and driving the saltwater wedge seaward or decreasing aquifer salinisation through blending with lower salinity surface water, such as treated wastewater (e.g., Masciopinto, Reference Masciopinto2013; Bachtouli and Comte, Reference Bachtouli and Comte2019; Alam et al., Reference Alam, Borthakur, Ravi, Gebremichael and Mohanty2021). For example, in Jakarta, Indonesia the effects of land subsidence and seawater intrusion are combatted by increasing groundwater recharge through the construction of water traps and drainage reservoirs and increasing vegetation cover (Pramono, Reference Pramono2021). However, if not managed properly MAR can introduce pollutants and pathogens into groundwater resources (e.g., Raicy et al., Reference Raicy, Parimala Renganayaki, Brindha, Elango, Elango, Goyal and Thomas2012; Casanova et al., Reference Casanova, Devau, Pettenati, Jakeman, Barreteau, Hunt, Rinaudo and Ross2016; Alam et al., Reference Alam, Borthakur, Ravi, Gebremichael and Mohanty2021). For example, a review of studies using stormwater for MAR revealed that dissolved organic carbon, selected metals and E. coli are effectively removed from recharge water during the infiltration process, but that trace organics and Enterococcus bacteria were still present (Alam et al., Reference Alam, Borthakur, Ravi, Gebremichael and Mohanty2021).

Policy-driven groundwater demand management has been successful in the drought-prone Central Coast region of California (USA) where groundwater supplies 90% of the drinking water (Langridge and Van Schmidt, Reference Langridge and Van Schmidt2020). In response to the severe drought of 2012–2016 experienced in the area, a Sustainable Groundwater Management Act (SGMA) was passed to mitigate groundwater storage loss (Langridge and Van Schmidt, Reference Langridge and Van Schmidt2020). Although the SGMA does not take into account groundwater losses prior to the passing of the Act, mitigation strategies related to groundwater include using flood flows rather than groundwater for MAR and irrigation purposes and developing local groundwater drought reserves to avoid irreversible groundwater losses during drought periods. The latter is achieved through, for example, using agricultural return flow for recharge, of which a portion is then available to farmers for irrigation during droughts (Langridge and Van Schmidt, Reference Langridge and Van Schmidt2020).

Another example of a coastal urban area that is prone to severe droughts and is predicted to be impacted by increased temperatures and decreases in surface water resources is that of Perth, Australia (population of 1.6 million) (Serrao-Neumann et al., Reference Serrao-Neumann, Renouf, Kenway and Choy2017). During the Millennium Drought (1996–2010) in southern Australia, it was proposed that the water supply in Perth be further augmented from the West Yarragadee aquifer, however, desalination of seawater was implemented instead (Lam et al., Reference Lam, Lant, O’Brien and Kenway2016). During 2013–2014 about 42% of Perth’s water supply was from groundwater, while desalination provided 39% of supply. More recently, MAR is also being considered to supplement water supply. Interestingly, the urban water use in Perth remained lower post-drought than that of pre-drought use, despite increased urban population, suggesting the value of social awareness and behavioural shifts in managing water scarcity. However, the energy use of the water supply system has doubled (Lam et al., Reference Lam, Lant, O’Brien and Kenway2016).

Several other interventions have been successful in decreasing/managing pollution and salinisation risks such as alternating/seasonal pumping regimes, pumping limits and hard-engineering approaches (e.g., subsurface dams, cut-off walls, semi-pervious subsurface barriers; Chang et al., Reference Chang, Zheng, Zheng, Zhang, Sun and Walther2019), rehabilitation of dune systems and the removal of alien trees (Giambastiani et al., Reference Giambastiani, Colombani, Greggio, Antonellini and Mastrocicco2017), and identifying sources of groundwater-borne pollutants (Michael et al., Reference Michael, Post, Wilson and Werner2017). Some of these interventions are not without adverse effects. For example, sub-surface dams can result in the accumulation of landward pollutants behind the barrier, increase inland soil salinisation and prevent groundwater discharge to the coastal zone (Chang et al., Reference Chang, Zheng, Zheng, Zhang, Sun and Walther2019).

Depending on the scale of a drought event, groundwater demand may be effectively managed by a range of options in the short to medium term (Parisi et al., Reference Parisi, Monno and Fidelibus2018), including alternative sources such as RWH, the recycling of water and desalinisation. RWH is perhaps the most universally applied method (Wurthmann, Reference Wurthmann2019) and can be used at a household (e.g., tank) or catchment (impoundment or for groundwater recharge) scale. For example, in Bangladesh, rainfall in coastal areas is higher than inland and RWH in urban areas is considered a suitable resource for drinking and residential purposes (Islam et al., Reference Islam, Afrin, Redwan and Rahman2015). Conversely in Florida USA, RWH used for residential outdoor irrigation could meet more than half the future water demand under a high population growth scenario and may perform better than other alternative sources such as reclaimed water (Wurthmann, Reference Wurthmann2019). The long-term reliability of RWH is, however, a concern since harvesting units are usually designed for current precipitation patterns and future rainfall variability is imminent due to climate change (Islam et al., Reference Islam, Afrin, Redwan and Rahman2015). In addition, at a catchment scale RWH through impoundments could reduce streamflow (Glendenning and Vervoort, Reference Glendenning and Vervoort2011).

The recycling of storm- and/or wastewater for irrigation purposes (Lavrnić et al., Reference Lavrnić, Zapater-Pereyra and Mancini2017), MAR (Alam et al., Reference Alam, Borthakur, Ravi, Gebremichael and Mohanty2021 and references therein) and even potable use (Lee and Tan, Reference Lee and Tan2016) is a climate-resilient water source. However, integration into water supply systems is not as readily accepted due to public perceptions as well as contamination and health risks (Gu et al., Reference Gu, Chen, Pody, Cheng, Zheng and Zhang2015; Garcia-Cuerva et al., Reference Garcia-Cuerva, Berglund and Binder2016). The use of reclaimed water for irrigation is considered a suitable option since the water already contains nutrients and therefore reduced fertiliser application is required (Lavrnić et al., Reference Lavrnić, Zapater-Pereyra and Mancini2017). Other innovations such as altered agricultural water supply systems (e.g., drip irrigation) or drought-resistant crops will be essential for future agricultural practices (Oude Essink, Reference Oude Essink2001; Lavrnić et al., Reference Lavrnić, Zapater-Pereyra and Mancini2017). Finally, desalination of brackish and seawater is a viable option for coastal urban water supply but is often expensive, energy-intensive and can have negative impacts on coastal biodiversity through brine by-products (Singh et al., Reference Singh, Jain, Sukhwani and Shaw2021 and see below case study).

Other management solutions include aquifer protection, through for example financial incentives to reduce groundwater abstraction and associated negative impacts. For instance, groundwater abstraction tax has been implemented in Jakarta and it was suggested that groundwater abstraction is spatially controlled (e.g., limiting abstraction to volcanic lithologies rather than subsidence-prone sedimentary lithologies) (Taftazani et al., Reference Taftazani, Kazama and Takizawa2022).

Drought management successes have also been achieved through adjusted water tariffs, integrated management of water resources, conservative water-use practices (residents and tourists), as well as the re-definition of management and policy systems (Parisi et al., Reference Parisi, Monno and Fidelibus2018; Singh et al., Reference Singh, Jain, Sukhwani and Shaw2021 and the references therein). Furthermore, it is crucial that groundwater modelling of water supply systems is included into drought policies and long-term planning. These models must take into account the present status of the aquifer(s) and be informed by baseline monitoring assessments and extraction rate information. However, a lack of monitoring and unmetered groundwater use poses a major challenge for groundwater management modelling (Hunter et al., Reference Hunter, Brooking, Reading and Vink2016; Keir et al., Reference Keir, Bulovic and McIntyre2019; Rochford et al., Reference Rochford, Ordens, Bulovic and McIntyre2022). Recent innovations in groundwater modelling have meant that hydrological deficits can be distinguished from the effects of over-abstraction and that aquifer responses can be linked to decision-making (Petersen-Perlman et al., Reference Petersen-Perlman, Aguilar-Barajas and Megdal2022). In addition, advances have been made in the understanding of aquifer vulnerability in terms of both volume and quality during droughts related to pumping and pollution. However, the effects of augmentation strategies such as MAR need to be incorporated more sufficiently (Petersen-Perlman et al., Reference Petersen-Perlman, Aguilar-Barajas and Megdal2022). Finally, it is essential that issues related to groundwater use during droughts are effectively communicated and that local stakeholders participate in managing groundwater resources (Petersen-Perlman et al., Reference Petersen-Perlman, Aguilar-Barajas and Megdal2022).

Both land and water management need to be proactive rather than reactive to ensure that coastal groundwater resources are managed adequately for the economy, health and environment (Michael et al., Reference Michael, Post, Wilson and Werner2017). Integrated coastal groundwater management needs to take into account geological, hydrological and biogeochemical complexity whilst also accounting for human complexity in terms of economic, cultural and decision-making factors (Michael et al., Reference Michael, Post, Wilson and Werner2017). Thus, environmental and hydrological connections between cities and the surrounding areas, future uncertainties related to water availability, and a holistic landscape view is required in planning (Serrao-Neumann et al., Reference Serrao-Neumann, Renouf, Kenway and Choy2017; see Figure 2).