Few factors are as important in determining water quality as land use/land cover (LULC). Many land use activities, including agriculture, urban development, mining, and commercial forestry, tend to be sources of diffuse pollution. By contrast, indigenous vegetation can act as a sink, thus providing some protection from diffuse anthropogenic contamination. Notwithstanding the large body of research demonstrating these facts, decision-makers require clear and accessible information to assist them in developing effective management plans that are fully cognisant of the manifold impacts of LULC on water resources. Reviewing the available literature, this article offers a critical overview of the typical impacts of LULC on water quality. An important strategy for managing water quality highlighted in this article is the maintenance of a sufficient amount of unfragmented natural vegetation, not only within riparian zones but also across catchment areas. However, knowledge gaps identified in this review indicate that further context-specific research is required to determine not only the types and minimum amount of vegetative cover required to protect water resources from diffuse pollution but also the potential impact of landscape fragmentation on the ability of natural vegetation to protect water resources. A critical discussion of these factors is therefore provided.

  • Improper land use management can have profound impacts on water quality.

  • While most anthropogenic land use activities generate diffuse pollution, natural vegetation can protect water resources by acting as a sink and biofilter.

  • Maintaining sufficient areas of unfragmented natural vegetation in both riparian areas and across whole catchments is an important water quality management strategy.

  • Local research is essential for determining not only which types of vegetation will offer water resources the most efficacious protection, but also for estimating the extent required to offer adequate levels of protection.

Anthropogenic land use/land cover (LULC) transformations are recognised as a force of global concern, causing environmental degradation across various ecosystems (Gomes et al. 2021; Taylor & Rising 2021; Henderson & Christian 2022; Liu et al. 2022; Richardson et al. 2023). In terms of aquatic ecosystem health and surface water quality, it seems almost impossible to overstate the significance of LULC within the surrounding landscape. There is, for instance, a large and growing body of research that establishes, through the use of robust statistical approaches and complex process-based models, that LULC has an unequivocal and decisive influence on water quality (Giri & Qiu 2016; Lintern et al. 2018; Ullah et al. 2018; Yuan et al. 2020; Bohenek & Sulliván 2022; Cheng et al. 2022; Goodspeed et al. 2022; Aalipour et al. 2023; Mashala et al. 2023; Qiu et al. 2023; Saeidi et al. 2023; Yao et al. 2023). It was with some prescience, therefore, that limnologist H.B.N. Hynes observed in one early article that ‘in every respect, the valley rules the stream’ (Hynes 1975, p. 12). Wear et al. (1998, pp. 619, 627) have since argued that the use and condition of land within catchment areas is ‘clearly one of the most important factors determining water quality.’ Therefore, according to Harris (2002, p. 343), rivers and other water bodies are ‘the ultimate integrators of our land use decisions’ (see also Falkenmark 2011, p. 13). However, some authors have cautioned that, in practice, the impacts of land use have not been given due regard in the formulation of plans and policies for the management and protection of water resources (Bandaragoda 2006; Falkenmark et al. 2014; Duda 2017). If this is the case, it is not for a lack of published research. Rather, it may be argued that the available research is either unintelligible or inaccessible to stakeholders and policymakers, who need to be provided with clear, relevant, and credible information that can guide their decision-making efforts (McDonnell 2008; Gooch & Stålnacke 2010; Halbe et al. 2013; Grigg 2021). This article, therefore, draws from a wide range of literature – including reference texts, academic reviews, and original research – to provide a concise contemporary overview of the typical impacts of different classes of LULC on water quality. To undertake a review of this literature, potentially relevant publications were initially identified by conducting searches on the Scopus, Google Scholar, and Web of Science databases. The results of these searches were further refined by manually reviewing document titles and abstracts for relevance. As the review progressed, citations contained in these articles were checked to locate additional relevant material that had not been identified in the original searches.

Based on this review, the present article highlights the specific negative impacts that agriculture, urban development, mining operations, and commercial forestry may have on water resources. It also notes the ameliorative effect that natural vegetation can have by intercepting and filtering contaminated surface runoff. It thus concludes that the maintenance and/or restoration of natural vegetation cover across catchments and within riparian areas is one of several important integrated management strategies that can help protect water resources from diffuse pollution. However, while this is a well-established fact, there are several important considerations discussed in this article – including issues related to scale and landscape configuration – that must guide the implementation of such a strategy. Consequently, local research is invaluable for informing the development of specific integrated management guidelines, such as land use thresholds, which may be used to protect water resources. After providing an overview of the impacts of anthropogenic land use activities on water quality, the remainder of this article focusses on a discussion of the role and importance of natural vegetation in water quality management and provides a critical discussion of some important caveats that require further consideration and/or investigation.

Although the terms ‘land use’ and ‘land cover’ are related and often used interchangeably, a distinction can be made between the two (Horning et al. 2010; Parece & Campbell 2015; Bohenek & Sulliván 2022). Strictly speaking, land cover refers only to the natural (biotic) and artificial (abiotic) features that cover the Earth's immediate surface in broad categories that may, for example, include built-up land, cropland, forested land, or water (Anderson et al. 1976; Thompson 1996; Schulze 2000; Thenkabail 2015; Heidkamp & Christian 2022). According to Giri (2012, pp. 9, 226), land cover refers to the ‘actual vegetative, structural, or other surface cover resulting from a given land use’ and thus represents the visible evidence of the latter. By contrast, land use refers to the human activities associated with a particular area of land in terms of utilisation, occupation, and/or management (Anderson et al. 1976; Thompson 1996; Giri 2012; Thenkabail 2015; Heidkamp & Christian 2022). It specifically refers to the manner in which the biophysical attributes of land are manipulated, managed, and exploited by humans (Schulze 2000, p. 13). Jansen & Gregorio (2002, p. 98) have therefore defined land use as ‘the arrangements, activities and inputs people undertake in a certain land cover type to produce, change or maintain it’ (see also Horning et al. 2010).

Concerning the relationship between LULC and water quality, Dabrowski et al. (2013, p. viii) have argued that ‘a distinction [can be] made between land use, associated with management practices influencing water quality, and land cover, which can be observed and mapped through earth observation technologies.’ The authors further observe that although it may be difficult to determine the details of specific land use practices (e.g., the amount or types of fertilisers applied to agricultural land) from remotely sensed earth observation data (e.g., satellite imagery or thematic land cover maps), land cover data may nevertheless serve as a general proxy for land use (ibid.). In any case, both land use (what people do on/with the land) and land cover (the land surface itself) influence water quality by determining not only the types and amounts of contaminants available for transport into receiving water bodies, but also how readily and by what pathways these contaminants may be mobilised and transported (Saeidi et al. 2023). As such, the acronym LULC (i.e., land use/land cover) is frequently used in the literature (as well as in this article) to encompass both.

Recent estimations suggest that humans have already transformed three-quarters of the Earth's surface, driven largely by expansions in urban and agricultural land use, and associated with significant losses of natural land cover in many regions (Song et al. 2018; Winkler et al. 2021). Moreover, despite net gains in forested land in some regions, it is predicted that urban and agricultural LULC will continue to expand into the future, with consequent impacts on many natural ecosystems (Nelson et al. 2010; van Asselen & Verburg 2013; Richardson et al. 2023). As the following sections demonstrate, these changes are likely to have significant impacts on freshwater systems and will therefore require careful integrated management and strategic planning that focusses on the maintenance and/or restoration of natural land cover in river catchments.

The impact of LULC on surface water quality is principally as a source of diffuse pollution (Chapman et al. 2020; Kajitvichyanukul & D'Arcy 2022). Concomitantly, Xu et al. (2023b) have described surface waters as the primary sink of pollutants in terrestrial ecosystems. It is this source/sink relationship, according to, D'Arcy et al. (2022a, p. 14), that inextricably links land and water resources. Accordingly, there are a variety of complex pathways and processes by which diffuse pollution enters water bodies from the land surface (Kajitvichyanukul & D'Arcy 2022; Zhang et al. 2023b). For instance, contaminants may be naturally present on or within the land surface (as in the case of naturally occurring mineral salts and nutrients in the soil) or later applied by humans (as in the application of fertilisers or pesticides). Various other land use activities may then expose or mobilise these pollutants, which are eventually transported during precipitation events into receiving water bodies through a variety of pathways (Lintern et al. 2018). Classes of LULC that are typically considered sources of diffuse pollution include urban, industrial, and other built-up areas, agricultural land, mines, and commercial forestry plantations (Dallas & Day 2004; Bosman & Kidd 2009; Falkenmark 2011; Chapman et al. 2020). As described in the following sections, each of these land uses is typically associated with a specific suite of pollutants and associated water quality impacts (Dabrowski et al. 2013; de Mello et al. 2020; Mirzaei et al. 2020; D'Arcy et al. 2022b).

Furthermore, landscape characteristics – especially the extent of impervious surfaces, the presence or absence of vegetation, and the spatial configuration of land cover within the landscape – will additionally influence the hydrological response of a catchment (Falkenmark et al. 1999; Ding et al. 2016; Aalipour et al. 2022). Higher proportions of impervious land cover, for example, will result in a ‘flashier’ hydrological response to precipitation events (i.e., reduced infiltration, shorter surface residence time, increased overland flow rates, shorter lag-times, reduced baseflow, and increased peak flow) (Schueler 1994; Paul & Meyer 2001; CWP 2003). This has associated impacts on water chemistry and pollution rates. Increased overland flow, for instance, flushes contaminants that have accumulated on impervious surfaces directly into receiving water bodies, resulting in higher pollutant loads (Day & Dallas 2011; Chapman et al. 2020). Initially, this may be offset, to some degree, by increased streamflow in the receiving water body, resulting in greater dilution of the received contaminants. However, in the longer term, reduced baseflow typical of streams in the urban environment will increase the relative concentration of pollutants in the water (Nobre et al. 2020; Li et al. 2022; Deng et al. 2023; Roldán-Arias et al. 2023).

Moreover, the spatial arrangement (i.e., configuration) of land cover within a landscape may influence its relationship with water quality by modifying, among other hydrological processes, the nature, pathways, and quantity of surface runoff generated during precipitation events (which will, in turn, influence the contaminant loads mobilised and transported into receiving water bodies from the land surface) (Shen et al. 2015; Song et al. 2021; Aalipour et al. 2023; Mo et al. 2023). It is frequently assumed, for example, that as natural vegetation cover becomes more fragmented, its ability to intercept, assimilate, and thus protect water resources from diffuse pollution will be reduced (Yirigui et al. 2019; Cole et al. 2020; Thomas et al. 2020; Fernandes et al. 2021; de Mello et al. 2022; Bowes et al. 2023; Zhang et al. 2023a). It is also commonly assumed that LULC directly adjacent to rivers and streams (i.e., riparian land use) is likely to have a more significant impact on water quality than land use further afield (Gove et al. 2001; Waite et al. 2010; Miller et al. 2011; Ou et al. 2016; Ramião et al. 2020; Han et al. 2023; Roldán-Arias et al. 2023; Xu et al. 2023b).

The following paragraphs describe the particular water quality impacts commonly associated with agricultural land, urban/built-up areas, mining operations, and commercial forestry plantations. A tabular summary of the major water quality parameters and pollutants derived from and/or influenced by these categories of LULC, as well as their typical impacts on aquatic ecosystems, is available in the Supplementary Material.

Agricultural land

There is an expansive body of literature documenting the manifold impacts of agricultural activities on water quality. These impacts are said to be considerable, well-established, and a major concern worldwide (Haygarth & Jarvis 2002; Dallas & Day 2004; Day & Dallas 2011; Mateo-Sagasta et al. 2018; Chapman et al. 2020). Runoff from agricultural land has been described by Dallas & Day (2004, p. 123) as a ‘complex effluent’ that typically contains elevated levels of nutrients, organic matter, dissolved salts, pesticides, sediments, and bacteria (see also Fraser 2002; Vigil 2003; Mateo-Sagasta et al. 2018; Chapman et al. 2020; Kronvang et al. 2020).

Perhaps the most widely reported impact of agriculture on water quality is the eutrophication of rivers and lakes due to the excessive input of nutrients (nitrogen and phosphorous) which may be applied to fields in the form of inorganic fertilisers or deposited in the excreta of grazing livestock (Haygarth & Jarvis 2002; Baker 2005; Meybeck et al. 2005; Mateo-Sagasta et al. 2018; Bohenek & Sulliván 2022). Nitrogen, when applied in excess of plant requirements, is highly mobile and easily leached from agricultural land by irrigation and/or precipitation. When land is cleared for planting, nitrogen loss increases as there are no plants to take up what is available in the soil (Haygarth & Jarvis 2002; Bohenek & Sulliván 2022). Phosphorous, while more resistant to leaching than nitrogen, is commonly transported along with the sediment particles to which it is often adsorbed, and even small amounts of phosphorous can have a significant impact on water quality (Haygarth & Jarvis 2002; Chapman et al. 2020). The resulting nutrient enrichment promotes excessive algal and macrophyte growth. This can cause a depletion of dissolved oxygen as these organisms respire, as well as when dead plant matter is decomposed by bacteria (Peters et al. 2005). Furthermore, blooms of cyanobacteria – a common symptom of eutrophication – can be toxic (Bohenek & Sulliván 2022).

Another widely reported water quality impact associated with agricultural land (particularly when irrigated) is salinisation (Dallas & Day 2004; Peters et al. 2005; Day & Dallas 2011; Bosman et al. 2018; Mateo-Sagasta et al. 2018; Chapman et al. 2020). Although irrigation water may not contain high concentrations of dissolved salts, any solutes it does contain will remain in the soil when it evaporates. Over time, these salts build up in the soil profile and are eventually leached into nearby water bodies when there is sufficient runoff (Dallas & Day 2004; Wepener et al. 2018). Additionally, irrigation tends to cause an increase in the water table, which draws dissolved minerals into the surface layers of the soil profile. These similarly accumulate in the soil as water evaporates and are eventually washed into receiving water bodies during precipitation events (Dallas & Day 2004). The clearing of land for cultivation may also increase groundwater recharge and raise water tables, with similar results (ibid.).

Turbidity and sedimentation are additional water quality problems frequently associated with agricultural land (Haygarth & Jarvis 2002; Dallas & Day 2004; Foley et al. 2005; Peters et al. 2005; Bosman et al. 2018; Mateo-Sagasta et al. 2018; Chapman et al. 2020). Sediments originate from land disturbed by cultivation, as well as from cleared/harvested fields where there is little or no vegetative cover to prevent erosion. Livestock may also promote erosion and the mobilisation of sediments through overgrazing and hoof action. Sediments, apart from their primary physical impact on aquatic environments (for which, see the Supplementary Material), may also transport other contaminants that become affixed to soil particles (including phosphorous, pathogens, and pesticides) (Day & Dallas 2011).

The application of toxic agrochemicals (including pesticides, herbicides, and fungicides) may further affect aquatic ecosystems (Haygarth & Jarvis 2002; Dallas & Day 2004; Foley et al. 2005; Bosman et al. 2018; Mateo-Sagasta et al. 2018; Chapman et al. 2020). These may enter water bodies through atmospheric deposition when they are applied to crops (i.e., ‘spray drift’), via accidental spills or through improper disposal, or when pesticide residues are flushed by precipitation or irrigation from the crops and/or fields to which they have been applied (Haygarth & Jarvis 2002; Dallas & Day 2004; Bosman & Kidd 2009).

Organic and bacteriological contamination, derived primarily from animal waste, are two other common impacts of agriculture on water quality (Haygarth & Jarvis 2002; Dallas & Day 2004; Meybeck et al. 2005; Peters et al. 2005; Mateo-Sagasta et al. 2018; Bohenek & Sulliván 2022). When organic matter in the water decomposes, potentially hypoxic conditions can develop as oxygen demand increases. Some bacterial species are also pathogenic and can cause disease in animals and humans. In some instances, animal waste may also contain traces of heavy metals which are added to animal feeds or given as food supplements (Haygarth & Jarvis 2002; Mateo-Sagasta et al. 2018). Antibiotics, hormones, and steroids (so called emerging pollutants or contaminants of emerging concern) may also be present in the excreta of treated livestock (Peters et al. 2005; Day & Dallas 2011; Mateo-Sagasta et al. 2018; Chapman et al. 2020; Bohenek & Sulliván 2022). Finally, reductions in flow from agricultural withdrawals can further influence water quality by reducing baseflow, thus increasing the in-stream concentrations of contaminants (Bosman et al. 2018; Chapman et al. 2020).

Urban/built-up land

Urban/built-up land, which often has a disproportionately negative impact on water quality relative to the area it occupies, is typically characterised by impervious surfaces and may include residential areas, informal settlements, commercial property, and industrial zones (Paul & Meyer 2001; CWP 2003; Dallas & Day 2004; Chapman et al. 2020). It is frequently reported, and with remarkable consistency, that water quality degradation typically occurs when urban/impervious cover in a catchment reaches between 10 and 15% (e.g., Schueler 1994; Arnold & Gibbons 1996; Paul & Meyer 2001; Brabec et al. 2002; Groffman et al. 2006; Tayyebi et al. 2015; Medupin et al. 2020; Song et al. 2020). Although urban land is often associated with point-source pollution (such as end-of-pipe discharges from wastewater treatment plants or industrial works), built-up areas can also be a significant source of diffuse pollution, which is arguably more difficult to regulate than discrete point-source discharges (Loague & Corwin 2005; Peters et al. 2005; Crooks et al. 2021; Day & Davies 2023). The wide range of contaminants found in urban runoff tend to accumulate on built-up surfaces through various deposition mechanisms and are later flushed into receiving water bodies during precipitation events (Peters & Meybeck 2000; Dallas & Day 2004; Chapman et al. 2020; Goodspeed et al. 2022). Diffuse urban pollution may also originate from leaks in sewerage systems or wastewater treatment works, as well as from seepages or spills at industrial sites (CWP 2003; Vigil 2003; Chapman et al. 2020; D'Arcy et al. 2022a). Leachate from landfills is another source of diffuse urban pollution (Bosman et al. 2018; Bohenek & Sulliván 2022). Owing to this diverse range of sources, urban stormwater usually contains elevated levels of most contaminants (Paul & Meyer 2001; CWP 2003; Ahmed et al. 2022).

Elevated concentrations of nutrients, which can originate from a variety of urban sources, are often found in runoff derived from built-up areas (Paul & Meyer 2001; CWP 2003; Meybeck et al. 2005; Chapman et al. 2020; Bohenek & Sulliván 2022). Phosphorous contributions of urban areas are even said to rival those of agricultural land in some cases (Paul & Meyer 2001, p. 342). High suspended solid and sediment loads are also common in urban runoff, and are derived especially from construction activities or the detritus that collects on roads and parking areas (CWP 2003; Vigil 2003; Dallas & Day 2004; Day & Dallas 2011). In addition, elevated levels of dissolved salts are frequently found in urban stormwater (Paul & Meyer 2001; Vigil 2003; Dabrowski et al. 2013; Bohenek & Sulliván 2022). Heavy metals, hydrocarbons, and a ‘whole suite’ of other persistent organic compounds are likewise typically present in urban runoff (Paul & Meyer 2001, p. 345; CWP 2003; Vigil 2003; Meybeck et al. 2005; Dabrowski et al. 2013; Chapman et al. 2020). Moreover, although pesticides are commonly associated with agricultural activities, they are also used in urban and industrial settings and thus often present in urban stormwater runoff (Paul & Meyer 2001; Gevao & Jones 2002; CWP 2003; Meybeck et al. 2005; Dabrowski et al. 2013; Chapman et al. 2020; Bohenek & Sulliván 2022). Bacterial contamination is also common in urban stormwater, especially in effluents derived from informal settlements where sanitation services are limited (Paul & Meyer 2001; CWP 2003; Dallas & Day 2004; Meybeck et al. 2005; Day & Dallas 2011; Dabrowski et al. 2013; Chapman et al. 2020; Bohenek & Sulliván 2022). Acid rain, a consequence of the atmospheric pollution often associated with urban and industrial areas, is another water quality concern linked to urban land use (Bosman et al. 2018; Chapman et al. 2020). Finally, urban runoff may also include a range of so-called emerging pollutants, including pharmaceuticals, hormones, solvents, and microplastics (Meybeck et al. 2005; Chapman et al. 2020; Bohenek & Sulliván 2022). The combined impacts of these contaminants include salinisation, acidification, turbidity, eutrophication, hypoxic conditions, and general toxicity. In addition, due to the increased proportion of impervious surfaces (which reduce permeability and increase overland flow rates), the total load of harmful material flushed into urban streams, especially during heavy rainfall events, is much increased (Paul & Meyer 2001; CWP 2003; Peters et al. 2005; Day & Dallas 2011; Parece & Campbell 2015; Chapman et al. 2020). This is compounded by generally reduced base-flows, thus elevating the in-stream concentration of these contaminants (Bohenek & Sulliván 2022). Unsurprisingly then, urban streams are some of the most seriously impacted in the world (Paul & Meyer 2001).

Mining operations

Pollution from mining operations is a worldwide problem that can have a significant impact on surface water quality (Dallas & Day 2004; Dabrowski et al. 2013; García et al. 2016; Chapman et al. 2020; de Mello et al. 2020). While working mines are undoubtedly a source of pollution, Chapman et al. (2020, p. 129) argue that decommissioned or abandoned mines tend to have a greater impact on water quality. This is due to the cessation of pumping operations and the resurgence and ingress of groundwater, which floods the exposed mine workings and becomes contaminated. When this contaminated water reaches the surface, it may flow into nearby water bodies. Leaks and spills from tailings dams are another common source of sediment and other contaminants (Peters et al. 2005). The nature of the effluent derived from mining activities logically depends on the type of mine, as well as the underlying geology of the area being mined. Ore-extracting substances such as cyanide, for example, may be present in runoff derived from gold mines (Dallas & Day 2004), whereas elevated concentrations of sulphate are a well-known impact of coal mining (Dabrowski et al. 2013). However, waters impacted by mining typically have low pH, high levels of suspended solids, and contain a variety of heavy metals (Dallas & Day 2004; Meybeck et al. 2005; Day & Dallas 2011; Bosman et al. 2018; Chapman et al. 2020). This leads to an increase in acidity, turbidity, sedimentation, and heavy metal concentrations in receiving waters. Contaminants may also precipitate out as a yellow or orange-brown flocculate (Chapman et al. 2020; Bohenek & Sulliván 2022). The most frequently documented water quality impact of mining is acid mine drainage (AMD) (Dallas & Day 2004; Weiner 2013; Bosman et al. 2018; Wepener et al. 2018; Boyd 2020; Chapman et al. 2020; Bohenek & Sulliván 2022). The very low pH levels that result from AMD are not only directly harmful to aquatic ecosystems but can also have additional impacts by increasing the availability and/or toxicity of other chemicals (a phenomenon known as synergism) (Dallas & Day 2004; Peters et al. 2005; Boyd 2020; Chapman et al. 2020; Bohenek & Sulliván 2022).

Commercial forestry

The impacts of commercial forestry operations on water quality are varied (Fulton & West 2002; Duffy et al. 2020). The disturbance of land by clearing, planting, and harvesting often results in increased erosion and the mobilisation of several contaminants, including eroded soil and sediments, nutrients, organic matter, and other debris (Vigil 2003; Dallas & Day 2004; Peters et al. 2005; Day & Dallas 2011; Dabrowski et al. 2013; Duffy et al. 2020). Increased nutrient loading is common, usually following the application of fertilisers to plantations, as well as through the leaching of soils disturbed during clearing, planting, or harvesting processes (Dallas & Day 2004; Peters et al. 2005; Day & Dallas 2011; Dabrowski et al. 2013; Duffy et al. 2020). The application of pesticides to commercial forestry plantations may also impact nearby receiving waters (Dabrowski et al. 2013). The clearing of stands can also result in an increase in the water table owing to reduced water uptake by plants. As this water evaporates from the surface soil layers, salts are left behind which may subsequently be flushed into streams when it rains. As such, common water quality impacts associated with commercial forestry plantations include turbidity and sedimentation, salinisation, nutrient loading, and pesticide toxicity.

Forestry operations also influence local hydrology and flow dynamics. Due to changes in the interception of precipitation, as well as in groundwater abstraction rates, streamflow may be reduced during afforestation and conversely increased during harvesting (Dallas & Day 2004). Plantations of non-indigenous trees can present a major threat to water resources by using more water than the natural vegetation they replace, reducing streamflow levels by up to 50% in some cases (Jewitt 2005). As indicated earlier, reduced flows may elevate the in-stream concentrations of many contaminants (Peters et al. 2005). Reduced interception due to clearing may also increase overland flow rates, thereby increasing the pollutant load of overland runoff. Notwithstanding the above, there is evidence which suggests that commercial forestry plantations, if managed properly, can offer beneficial ecosystem services – including buffering, erosion control, water quality maintenance, and flow regulation – and thereby have positive impacts on water quality in the long-term (Ide et al. 2019; Malherbe et al. 2019; Duffy et al. 2020). It would therefore appear that the impact of commercial forestry depends largely on factors such as the species grown and the manner in which plantations are managed (de Mello et al. 2020).

While most anthropogenic classes of LULC are sources of diffuse pollution (see above), areas of natural vegetation (e.g., native forests, grasslands, shrublands, and wetlands) act as sinks by filtering, assimilating, and transforming pollutants before they enter receiving water bodies (de Mello et al. 2017; Brogna et al. 2018; Bohenek & Sulliván 2022; Cheng et al. 2022; Feng et al. 2023; Qiu et al. 2023; Woznicki et al. 2023). Natural vegetation also plays an important role in regulating overland flow rates, thereby reducing erosion and the overall pollutant load of surface runoff (Malherbe et al. 2019; Yirigui et al. 2019; Sun et al. 2020; Caldwell et al. 2023). Riparian vegetation additionally helps to shade water, keep it cool, and so regulate dissolved oxygen concentrations (Vigil 2003). It is therefore widely acknowledged that while increased proportions of urban and agricultural land cover tend to be correlated with negative water quality impacts, undeveloped areas of natural vegetation are generally associated with improved water quality (de Mello et al. 2017, 2022; Brogna et al. 2018; Fernandes et al. 2021; Piffer et al. 2021; Wang et al. 2021; Cheng et al. 2022; Li et al. 2022; Qiu et al. 2023). The clearing of natural vegetation, and its transformation into other pollution-generating land uses, is therefore singularly detrimental to water quality (Harris 2002; Waite 2014; Parece & Campbell 2015; Brogna et al. 2017; Caldwell et al. 2023). This makes the preservation and/or restoration of natural vegetation within catchments and riparian zones an important water quality management strategy (Norris 1993; Haycock et al. 2001; Stutter et al. 2012; de Mello et al. 2017; Sun et al. 2020; Cheng et al. 2022; Cooke et al. 2022). Pengelly & Fishburn (2002, p. 409) conclude that ‘native vegetation is a valuable resource and an important key to maintaining the health of the land and waterways,’ while Piffer et al. (2021, p. 1) similarly suggest that ‘maintaining or restoring native vegetation cover is a promising intervention to sustain adequate water quality.’

A common strategy for protecting water quality from diffuse pollution is the establishment and maintenance of ‘buffer zones’ of riparian vegetation that can intercept overland flow before it reaches receiving waters (Norris 1993; Haycock et al. 2001; Stutter et al. 2012; Sweeney & Newbold 2014; Cole et al. 2020; Petersen et al. 2020). Cooke et al. (2022, p. 183) summarise: ‘From reducing erosion and flood damage to maintaining cool water temperatures, filtering pollutants, protecting critical habitats, and enabling lateral connectivity, intact riparian zones mitigate many of the threats that degrade freshwater ecosystems.’ While it is generally recommended that vegetated buffers should be at least 30 m wide, it is arguable that the greater the width of the buffer, the greater the benefit provided (Haycock et al. 2001; Sweeney & Newbold 2014). Nonetheless, while the maintenance and/or restoration of natural vegetation within catchment areas and riparian zones is a critical water quality management strategy, there are a number of important conceptual, methodological, and practical issues that require consideration. These are discussed in turn in the sections that follow.

Defining and classifying ‘natural vegetation’

In most of the studies in which the impacts of LULC on water quality have been investigated using remotely sensed land cover data, forests are typically assumed, whether implicitly or explicitly, to be representative of natural vegetation (e.g., Omernik et al. 1981; Maloney & Weller 2011; Yu et al. 2013; Piffer et al. 2021; Zhang et al. 2021; Allafta & Opp 2022; de Mello et al. 2022; Caldwell et al. 2023; Qiu et al. 2023). This assumption may be reflective of the fact that most published studies have been conducted in temperate regions of the United States and Europe, where forests are the dominant natural biome (Baker 2005; Tromboni & Dodds 2017; Kronvang et al. 2020; Bohenek & Sulliván 2022; Prakoso et al. 2023). Thus, many studies emphasise the need to maintain sufficient areas of forested land within catchments in order to protect water quality (e.g., Caldwell et al. 2023). However, in ecologically and climatically disparate regions (such as savannas, grasslands, and scrublands), the extent of forested land may not be the most appropriate land cover metric by which to estimate the condition of catchments. For this reason, perhaps, other studies have broadened their classification of natural land cover to include an aggregation of other classes of indigenous vegetation, including woodland, shrubland, grassland, and wetlands (Sponseller et al. 2001; Tiner 2004; Shiels 2010; Bierschenk et al. 2012; Pandey et al. 2012; Iñiguez-Armijos et al. 2014). However, while aggregating all locally-occurring types of natural vegetation into a single land cover class may seem more appropriate (not to mention convenient), Cole et al. (2020, p. 3) have noted that the morphological and functional traits of different plant species can influence their impact on water resources. Thus, depending on their physiological, structural, and life-cycle characteristics, different types of vegetation may be more or less effective at intercepting and removing contaminants from overland flow, and thus at protecting water resources from diffuse pollution (ibid.). For instance, while grassland is typically expected to act as a detention medium by trapping contaminants in overland flow before they reach receiving waters, there is evidence which suggests that grassland may not consistently serve as an effective buffer against diffuse pollution (Ahearn et al. 2005; Amiri & Nakane 2006; Xiao & Ji 2007; Ding et al. 2013; Shen et al. 2014; Chen et al. 2016, 2021; Vrebos et al. 2017; Lacher et al. 2019; Dymek et al. 2021; Zhou et al. 2022). Given these complexities, it is essential for researchers and practitioners to consider the most appropriate land cover metric by which the extent and condition of natural vegetation within catchment areas may assessed, and to do so on a case-by-case basis, taking both local ecological conditions and the specific aims/objectives of the study (or management strategy) into account.

The importance of a multiscale approach

While vegetated riparian buffer zones are indisputably important for maintaining water quality, the cumulative effects of LULC across the entire catchment may be too great to be effectively mitigated or offset by riparian land use (Allan 2004a, 2004b; Brabec 2009; Tran et al. 2010; Tromboni & Dodds 2017; Ramião et al. 2020; Thomas et al. 2020). Thus, despite the logical importance of managing LULC within riparian areas, land use decisions across the entire catchment may be equally as important – if not more important – as those made within riparian zones (Brabec et al. 2002; de Mello et al. 2020; Ramião et al. 2020). In support of this, several studies which have compared the relative influence of LULC at different scales have reported that LULC at the whole-catchment scale was more significant, in terms of water quality impacts, than riparian LULC (Sliva & Williams 2001; Sponseller et al. 2001; Buck et al. 2004; Magierowski et al. 2012; Nielsen et al. 2012; Pratt & Chang 2012; Ding et al. 2016; Clément et al. 2017; de Mello et al. 2018; Park & Lee 2020; Łaszewski et al. 2022; Deng et al. 2023). Several authors have thus emphasised the necessity of taking a multiscale approach when researching and managing the impacts of LULC on water quality (Strayer et al. 2003; Schiff & Benoit 2007; Zhou et al. 2012; Ding et al. 2016; de Mello et al. 2018; Park & Lee 2020; Song et al. 2021; Pei et al. 2023; Siqueira et al. 2023). According to Nobre et al. (2020, p. 7) for example, ‘the finding that land use in the buffer area, as well as in the catchment as a whole, can modulate water quality has implications for water management.’ As such, it would be inappropriate to focus on managing land use at one spatial scale (e.g., within riparian areas) while ignoring land use at other scales (e.g., across whole catchments) (see also Cole et al. 2020, p. 8).

Landscape configuration

While the proportion of a landscape occupied by different classes of LULC (i.e., landscape composition) will naturally have an impact on water quality, it has also been suggested (and subsequently demonstrated in several studies) that the spatial arrangement, position, and/or distribution of LULC (i.e., landscape configuration) may also significantly influence its impact on water quality (Lintern et al. 2018; Zhang et al. 2019; de Mello et al. 2020; Dymek et al. 2021; Aalipour et al. 2022; Cheng et al. 2022; Xu et al. 2023a). Specifically, as noted earlier, it is often assumed that as patches of natural vegetation cover become more fragmented (a typical consequence of human development) their ability to intercept and protect water resources from diffuse pollution is reduced. Several studies have investigated the influence of landscape configuration on water quality, and many have found that increased fragmentation, especially of natural vegetation such as forests, is associated with poorer water quality (Lee et al. 2009; Bateni et al. 2013; Shen et al. 2014; Ye et al. 2014; Ding et al. 2016; Liu & Yang 2018; Yirigui et al. 2019; Zhang et al. 2019; Liu et al. 2021; Wu & Lu 2021; de Mello et al. 2022; Zhong et al. 2022; Qiu et al. 2023). According to Lintern et al. (2018, p. 15), for example, ‘small and fragmented forests’ are ‘not as effective at reducing the contaminants contained in runoff from other sources (e.g., urban, agricultural land uses) within the catchment.’ Xu et al. (2023a, p. 10) have thus emphasised the necessity of maintaining ‘large and intact’ areas of forest to protect water resources. Logically, however, the scale at which fragmentation is likely to have the most significant effect on vegetation's ability to protect water resources from diffuse pollution is within riparian buffer zones, where natural vegetation functions as a ‘last line of defence’ for water resources (Song et al. 2021, p. 1).

Thresholds of natural vegetation

Several studies have demonstrated that aquatic ecosystems may experience abrupt and/or undesirable changes in response to transformations of their surrounding landscapes (Brabec et al. 2002; Allan 2004b; Dodds et al. 2010; Tromboni & Dodds 2017; D'Amario et al. 2019; Li et al. 2021; Zhong et al. 2022; Mo et al. 2023). Tayyebi et al. (2015, p. 103), for instance, have observed that ‘research has demonstrated that land use tipping points in water quality and aquatic health occur when a certain percentage of watershed or catchment exceeds a certain amount of urban and agricultural land use’ (see also Capon et al. 2015; Grimstead et al. 2018; Liu et al. 2021). Provided that they are accurately estimated, such thresholds may provide policymakers with important insights about, and objective targets for, the protection of water resources (Wang et al. 2023; Xu et al. 2023b).

While thresholds of urban and agricultural land are most frequently reported, Nash et al. (2009, pp. 358–359) have argued that ‘maintaining and preserving natural land cover by a certain level may help in providing healthier surface water quality’ (emphasis added). In addition, both Attua et al. (2014, p. 66) and Cecílio et al. (2019, p. 49) have argued that maintaining an ‘adequate’ level of natural vegetation in catchments can improve water quality. However, what ‘certain level’ of natural vegetation may be considered ‘adequate’ for the protection of water quality is a question that requires further study (Brabec et al. 2002; Death & Collier 2010; Iñiguez-Armijos et al. 2014; Hanna et al. 2021). The research conducted to date, for instance, seems to indicate that the answer is dependent on several contextual factors, with estimated thresholds ranging from 40 to 90% (likely owing to differences in the methods and metrics used, as well as local environmental factors). Table 1 lists examples of publications in which these findings have been reported.

Table 1

Publications in which thresholds of natural vegetation have been reported

PublicationStudy regionFinding
Black et al. (2004)  Pacific Northwest, United States Health of in-stream macroinvertebrate communities began to decline rapidly when forest cover across the catchment fell below 70–80% 
Sheeder & Evans (2007)  Pennsylvania, United States When considering nutrient concentrations and sediment loads, ‘unimpaired’ catchments had an average of 78% forest cover 
Death & Collier (2010)  Waikato Region, New Zealand Catchments with 80–90% indigenous vegetation cover were associated with fauna indicative of ‘clean’ water quality, while streams draining catchments with 40–60% vegetation cover retained 80% of the biodiversity found in pristine streams 
Iñiguez-Armijos et al. (2014)  Southern Andes, Ecuador The ecological condition and macroinvertebrate biodiversity of streams were found to be ‘good’ when the vegetative cover was above 70% 
Clément et al. (2017)  Eastern Canada In areas dominated by agriculture, eutrophication was prevalent in catchments with less than 47% forest cover 
Kändler et al. (2017)  River Nisa, Czech Republic and Germany Catchments with more than 70% forest cover tended to have the lowest concentrations of nutrients and heavy metals 
Pond et al. (2017)  West Virginia, United States When evaluating both physiochemical and biological indicators, impairment occurred when catchment-scale forest cover fell below 60% 
Ding et al. (2021)  Liaoning Province, China A significant decline in in-stream taxa abundance was observed when catchment vegetation fell below approximately 60% 
Zhong et al. (2022)  Dianchi Lake Basin, China A minimum of 45% natural vegetation cover was required to maintain nitrogen concentrations within acceptable levels 
PublicationStudy regionFinding
Black et al. (2004)  Pacific Northwest, United States Health of in-stream macroinvertebrate communities began to decline rapidly when forest cover across the catchment fell below 70–80% 
Sheeder & Evans (2007)  Pennsylvania, United States When considering nutrient concentrations and sediment loads, ‘unimpaired’ catchments had an average of 78% forest cover 
Death & Collier (2010)  Waikato Region, New Zealand Catchments with 80–90% indigenous vegetation cover were associated with fauna indicative of ‘clean’ water quality, while streams draining catchments with 40–60% vegetation cover retained 80% of the biodiversity found in pristine streams 
Iñiguez-Armijos et al. (2014)  Southern Andes, Ecuador The ecological condition and macroinvertebrate biodiversity of streams were found to be ‘good’ when the vegetative cover was above 70% 
Clément et al. (2017)  Eastern Canada In areas dominated by agriculture, eutrophication was prevalent in catchments with less than 47% forest cover 
Kändler et al. (2017)  River Nisa, Czech Republic and Germany Catchments with more than 70% forest cover tended to have the lowest concentrations of nutrients and heavy metals 
Pond et al. (2017)  West Virginia, United States When evaluating both physiochemical and biological indicators, impairment occurred when catchment-scale forest cover fell below 60% 
Ding et al. (2021)  Liaoning Province, China A significant decline in in-stream taxa abundance was observed when catchment vegetation fell below approximately 60% 
Zhong et al. (2022)  Dianchi Lake Basin, China A minimum of 45% natural vegetation cover was required to maintain nitrogen concentrations within acceptable levels 

A review of the literature published by the Open Space Institute concludes that water quality tends to deteriorate when the proportion of forested land falls below 60–90% of the total catchment area, while observing that a threshold of at least 70% forest cover appears to represent the consensus (Morse et al. 2018). However, the review acknowledges that threshold values will largely depend on other site-specific factors (including local ecology, climate, and geology), as well as on the nature and intensity of nearby anthropogenic land uses. This underlines the necessity of estimating thresholds by conducting localised studies that can account for the influence of these variables.

In addition, as the relationship between land use and water quality is complex, it is important to note that no single LULC metric, nor any threshold based on these metrics, can fully account for the potential impacts that LULC may have on water quality in a given landscape (Klein 1979; Brabec et al. 2002; Zampella et al. 2007; Brabec 2009; Dabrowski & de Klerk 2013; Shen et al. 2015). As such, thresholds of natural vegetation should not be regarded as a water quality management panacea. Xu et al. (2023a, p. 1), for instance, observe that, ‘due to the many causes and complicated processes causing water quality changes, controlling and preventing stream water quality degradation remains challenging.’ It must therefore be borne in mind that there are several other potential sources and/or causes of impairment that cannot be addressed through land use management alone. An integrated management approach, which considers other potential sources of diffuse pollution and strategies by which these may be mitigated, is thus essential.

From an integrated catchment management perspective, few factors are as important for determining the quality and condition of surface waters as LULC. Although direct ‘point-source’ inputs (such as end-of-pipe discharges from wastewater treatment works) can have major impacts on water quality if not properly regulated, it is often the diffuse pollution derived from anthropogenic land use transformations (including agriculture, urban development, mining, and commercial forestry) that has the most profound impact on water resources. Natural vegetation, by contrast, offers important ecosystem services by intercepting and remediating contaminated surface runoff before it reaches receiving waters. Acknowledging, therefore, that anthropogenic land uses tend to have a detrimental impact on water quality, and that the preservation and/or restoration of natural vegetation can help to mitigate these impacts, it follows that, in order to protect water resources, land use must be carefully managed by limiting potentially harmful land use transformations. This involves, as part of a broader integrated management strategy, ensuring that sufficient areas of unfragmented natural vegetation are maintained across catchment areas and within riparian zones. It is vital, therefore, for legislators and policymakers to require the protection and/or restoration of undisturbed natural areas as part of integrated catchment management plans. However, owing to the additional influence of local ecological and environmental conditions, determining the appropriate types and/or requisite amount of natural vegetation necessary to protect water resources in any given region requires context-specific local research that can inform such policy directives.

The author is grateful to the University of Cape Town for funding the doctoral study from which this article emanates.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.

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