Microbiological contamination

Approximately 43% of drinking water advisories (DWAs) in Indigenous communities are attributed to high levels of bacteria (e.g., E. coli and total coliforms) (Health Canada, 2010; Lane et al., Reference Lane, Trueman, Locsin and Gagnon2020). Although chlorine is added to water during disinfection processes to eliminate pathogens, the treated water can still be susceptible to contamination (e.g., autochthonous bacteria, protozoa and suspended solids) within the distribution system or transportation trucks. Adding 0.2 mg/L residual chlorine prevents the regrowth of harmful microorganisms and bacteria in the distribution system (EPA, 2021). The most frequent pathways of microbiological contamination in drinking water are polluted seepage through cracks in the broken pipes and cisterns, and biofilm formation on the inner walls of distribution systems (Bashar, Reference Bashar2021; Stefan et al., Reference Stefan, Bosomoiu and Teodorescu2023).

A study by Farenhorst quantified bacteria in drinking water sources of a fly-in First Nations community in Manitoba. No coliforms or E. coli were detected in the water treatment plant (WTP) tap, indicating that the treatment process was efficient at removing bacteria. However, total coliform levels increased significantly along the distribution and storage pathway, following the trend: buckets/drums in homes without running water > piped homes = water truck > source water = cistern > community pipe = water truck. The cistern and source water (lake) had similar E. coli levels (65–97 CFU/100 mL). In communities without running water, levels of E. coli in standpipes (1,260 CFU/100 mL) and buckets/drums in homes (7,780 CFU/100 mL) were dramatically higher than in lake water, indicating massive bacterial growth after treatment. E. coli concentration in the tap water of piped homes (1 CFU/100 mL) and water truck (6 CFU/100 mL) was considerably lower compared to homes without distribution systems (Farenhorst et al., Reference Farenhorst, Li, Jahan, Tun, Mi, Amarakoon, Kumar and Khafipour2017).

Water supply arrangements involving the transport of drinking water from WTPs and storage tanks are common in Manitoba SRR communities. It is reported that 50% of Indigenous communities on reserves in Manitoba receive piped water from WTPs. Approximately 31% are on cisterns, and their water quality is relatively affected by bacteria, depending on the type of cistern and the material (below-ground concrete and fiberglass and above-ground polyethylene insulated shelters) (Amarawansha et al., Reference Amarawansha, Zvomuya and Farenhorst2021). This study on three communities from the Tribal Councils of Swampy Cree, West Region and Island Lake showed that the tap water in households relying on cisterns exhibited a comparatively higher level of contamination with coliform bacteria and E. coli compared to tap water in households connected to piped water systems, especially in the late spring. Furthermore, it is recommended that below-ground concrete or fiberglass cisterns be replaced with above-ground cisterns or piped water to reduce the risk of waterborne illnesses associated with elevated bacterial levels.

Depending on the materials used in manufacturing, different bacterial colonies can grow in water storage tanks and cisterns. Corrosive materials in metal storage tanks facilitate the growth of bacteria such as Gallionella spp. and Desulfovibrio spp., leading to microbial-induced corrosion and biofilm development, which may introduce pathogens into the water supply (Rilstone et al., Reference Rilstone, Vignale, Craddock, Cushing, Filion and Champagne2021). While plastic cisterns are commonly used, they can release carcinogenic compounds due to polymer degradation, especially under high temperatures and chemical exposure. Additionally, poor maintenance, open storage and leaks allow contamination by Salmonella spp. and Campylobacter spp. from the surroundings (Zambrano-Alvarado and Uyaguari-Diaz, Reference Zambrano-Alvarado and Uyaguari-Diaz2024). Studies in fly-in First Nations communities located in Manitoba demonstrated E. coli levels as high as 60,000 CFU/100 mL, total coliform counts >1,000 CFU/100 mL and heterotrophic counts >1,000 CFU/100 mL in distribution systems (pipelines and fiberglass tanks) (Farenhorst et al., Reference Farenhorst, Li, Jahan, Tun, Mi, Amarakoon, Kumar and Khafipour2017; Zambrano-Alvarado and Uyaguari-Diaz, Reference Zambrano-Alvarado and Uyaguari-Diaz2024).

Two other studies conducted in 2016 and 2019 examined antibiotic resistance genes, free residual chlorine and fecal bacteria (E. coli and coliforms) in three Indigenous communities in the Island Lake Region of Manitoba. A total of 122 samples were taken from three main points of the drinking water distribution network: GW-to-well-to-house, GW/lake water-to WTP-to-pipeline-to house and GW/lake water-to WTP-to truck-to plastic/concrete cisterns-to house. Samples collected from tap and cisterns had lower total coliform and E. coli counts than lake water, but higher than those in WTP samples. Despite no detectable bacterial contamination in WTP output samples, high levels of E. coli, total coliforms and antibiotic resistance genes were found once water entered the distribution systems, indicating infrastructural failure. Consistent with previous findings, samples collected from plastic cisterns exhibited the highest contaminant levels and the lowest free chlorine residual (Fernando et al., Reference Fernando, Tun, Poole, Patidar, Li, Mi, Amarawansha, Dilantha Fernando, Khafipour, Farenhorst and Kumara2016; Mi et al., Reference Mi, Patidar, Farenhorst, Cai, Sepehri, Khafipour and Kumar2019).

In Alberta, 56 SWSs on First Nation reserves were evaluated from a public health risk standpoint. Fifty systems were reported as high risk, five as medium and one as low risk. A numerical ranking was conducted in accordance with the GCDWQ guidelines. The major challenge, in addition to cultural, social, political and economic circumstances, was the lack of source water characterization, especially on pathogenic microorganisms, such as protozoan parasites, cysts, fecal bacteria and Cryptosporidium (Smith et al., Reference Smith, Guest, Svrcek and Farahbakhsh2006).

Microbiological contamination, often by E. coli and total coliform bacteria, remains the main cause of BWAs in Indigenous communities. Surface water sources generally pose a higher risk of microbiological contamination compared to GW, due to direct exposure to runoff, wildlife, wastewater inputs, seasonal events and algal blooms. However, studies across Manitoba and Alberta show that microbiological risks originate along the treatment-to-tap pathway, rather than in treated water quality. Detected levels of E. coli and total coliforms in many posttreatment samples from distribution infrastructure, cisterns and transportation trucks exceeded GCDWQ as high as 60,000 CFU/100 mL regardless of source water type. Microbiological contamination of drinking water in small communities is linked to elevated prevalence of waterborne diseases, bacterial gastroenteritis and skin infections such as impetigo, emphasizing the public health importance of controlling posttreatment contamination in decentralized systems. (Bradford et al., Reference Bradford, Okpalauwaekwe, Waldner and Bharadwaj2016).

Heavy metals and trace minerals

HMs and trace minerals represent a critical category of contaminants in SRR communities’ source waters, given their persistence, widespread occurrence and potential health risks. Elements such as Fe, Zn, Cu and Mn are essential to the human body at very low levels, but become toxic at higher concentrations (Mehri, Reference Mehri2020). Radioactive metals such as Uranium (U) naturally occur and may be released through erosion and weathering of rocks and soils. Additionally, aesthetic features and physical parameters are affected by metal and mineral impurities, including Fe, Mn, Zn, sodium (Na), Pb, Cu, Sb and Al. Prolonged exposure to HMs is linked to a range of detrimental health issues and cancer development (Rehman et al., Reference Rehman, Fatima, Waheed and Akash2018).

Arsenic is one of the most dangerous contaminants in source water, particularly GW, causing at least 5% of long-term water advisories in First Nations communities (Rehman et al., Reference Rehman, Fatima, Waheed and Akash2018). As shown in Table 1, the MAC for As was recommended at 0.01 mg/L by GCDWQ. Private well owners are recommended to test their water every 2–5 years for As, which is a very long monitoring period (Chappells et al., Reference Chappells, Parker, Fernandez, Conrad, Drage, O’Toole, Campbell and Dummer2014).

Natural forms of As, arsenopyrite (FeAsS), loellingite (FeAs₂), orpiment (As₂S₃) and realgar (AsS) are formed because of volcanoes and eroded arsenic-bearing rocks. It is also widely used in human activities such as agriculture, industrial chemical manufacturers, mining, iron ore smelting, pulp and paper (Lum et al., Reference Lum, Schoepfer, Jamieson, McBeth, Radková, Walls and Lindsay2023). For instance, Rabbit Lake Uranium mine mill tailings in Saskatchewan, Canada, caused As distribution in the adjacent aquifers. Five monitored wells had high As concentrations ranging from 9.6 to 71 mg/L (Berthiaume, Reference Berthiaume2023; Donahue and Hendry, Reference Donahue and Hendry2003). High concentrations of As (26.8 mg/L) and Sb (0.92 mg/L) in surface water originated from Giant gold-mine residues found in Yellowknife, Northwest Territories, which impacted ponds, sediments, vegetation downstream and the source water of local Indigenous communities (Fawcett et al., Reference Fawcett, Jamieson, Nordstrom and McCleskey2015).

In Alberta, among 2,817 private domestic well samples, 62.2% were high in As (50% > 10 μg/L, 12.2% > 50 μg/L) (Moncur et al., Reference Moncur, Paktunc, Jean Birks, Ptacek, Welsh and Thibault2015). In rural areas, 48% of 42 GW samples exceeded 25 μg/L, with a maximum concentration of 5.34 mg/L (Government of Alberta, 2011). According to a report by the British Columbia Ministry of Environment, elevated As concentrations impacted the water quality of 14 aquifers across the province. Ninety-eight private wells were tested, where 43% > 10 μg/L (Wilson et al., Reference Wilson, Schreier, Brown, Abbotsford and Region2008). However, the report did not specify which aquifers were affected (BC Government, 2007).

In Manitoba, the concentration of As in 10,563 GW samples was investigated. Several similar studies were carried out in other regions. The results are as follows: Manitoba: 5.9% >10 μg/L, 25 μg/L < 2% < 62 μg/L (McGuigan et al., Reference McGuigan, Hamula, Huang, Gabos and Le2010), Newfoundland and Labrador: sample number N/A, from 70 wells, one well at 48–55 μg/L and another at 335–368 μg/L (NL Environment, 2009), Nova Scotia: sample number N/A, from 31 wells, 4 of them at 10-15 μg/L, one well at 58 μg/L (NS Environment, 2009), Québec: 10 wells >101 μg/L and Saskatchewan: 48 wells >10 μg/L were the highest recorded concentrations of As (McGuigan et al., Reference McGuigan, Hamula, Huang, Gabos and Le2010).

Cobalt and Crosswise lakes, located in Timiskaming First Nations lands, Ontario, were affected by As and other trace metals (silver, cobalt and nickel arsenides, sulfide-arsenides, sulpho-arsenites, antimonides and sulpho-antimonites and sulpho-bismuthites) as a result of abandoned silver mine sites and unregulated disposal of tailings and ore with high As content (Chen et al., Reference Chen, Yu and Belzile2019).

When aggregated across provinces, As is the most frequently investigated HM in SRR and Indigenous drinking water studies. Alberta and British Columbia report some of the highest As exceedances in private and community wells, whereas Manitoba and Atlantic Canada show lower exceedance frequencies but widespread detection. Ontario and British Columbia account for the largest share of samples in multi-province studies, indicating uneven geographic research coverage, with fewer systematic assessments in northern and remote regions.

Health Canada’s 2019 GCDWQ specified Mn as a neurotoxic trace mineral with a MAC of 0.120 mg/L. Previously, Mn was considered only as an aesthetic issue in water. However, it has been proven that childhood exposure to high levels of Mn can cause cognitive and attention deficits. Mn in water is occasionally associated with Pb, and co-exposure to Pb and Mn is harmful to human health (Singh et al., Reference Singh, Bist and Choudhary2024).

A comprehensive 2020 study in partnership with the Atlantic Policy Congress of First Nation Chiefs analyzed samples from 47 First Nations with small systems for Pb, As and Mn. The sampling process complied with annual frequency and regulatory guides (concentration of trace metal) over 12 years. Overall, 32 SWSs supplying 3,440 residents relied on GW, and 50 small systems supplying 660 residents relied on surface water. The communities were not under a municipal transfer agreement and had their water and wastewater services owned by First Nations. Otherwise, the municipality was responsible for providing those services. Of the 16,920 samples examined, 22, 18 and 11 systems exceeded Mn, Pb and As MAC, respectively. In other words, 96% of Pb, 93% of As and 84% of Mn in the analyzed samples had concentrations below the guideline. The maximum concentrations were reported at 47,000 μg/L for Mn, 3,000 μg/L for Pb and 250 μg/L for As. Although most samples did not exhibit high levels of contaminants, increasing the sampling frequency to determine contaminant patterns was recommended (Lane et al., Reference Lane, Trueman, Locsin and Gagnon2020).

A critical distinction emerging from the literature is that detection of metals does not necessarily imply regulatory non-compliance or health risk. While As, Mn and Pb are frequently detected across supply wells, most concentrations reported in large monitoring programs fall below GCDWQ guideline values. These results suggest that high HMs concentrations tend to be localized and system-specific rather than widespread, underscoring the importance of monitoring frequency and spatial distribution in risk assessment.

In a broader 2021 study, water samples were collected from households in 91 randomly selected First Nations south of the 60th parallel, including British Columbia, Alberta, Saskatchewan, Manitoba, Ontario and Quebec-Labrador over 10 years. All adult household members identified as First Nations agreed to participate in the sampling process, surveys and interviews. Depending on the source water used in a community, samples were taken from taps, pipeline ends, miscellaneous points within the system, wells, lakes, springs, trucked water and bottled water (in case of DWA). The results are summarized in Table 2. According to the findings, As levels exceeded the GCDWQ in some regions due to background concentrations and natural activities. However, it was not reported along with the HMs results. In total, 2,933 samples were collected. Ontario, with 22% of all samples and British Columbia, with 20%, were the most tested provinces (Schwartz et al., Reference Schwartz, Marushka, Chan, Batal, Sadik, Ing, Fediuk and Tikhonov2021).

Table 2.

Metals concentration in drinking water in participant First Nations by region (Schwartz et al., Reference Schwartz, Marushka, Chan, Batal, Sadik, Ing, Fediuk and Tikhonov2021)

Comparison of Table 2 with the GCDWQ thresholds in Table 1 indicates that Pb concentrations exceeded their respective MACs (0.005 mg/L) in several regions, most notably Ontario and Atlantic, reaching levels of potential health concern. In contrast, Fe, Mn, Al, Cu and Zn frequently exceeded aesthetic objectives or operational guidelines rather than health-based MACs, suggesting that their primary impacts relate to infrastructure fouling and system performance rather than direct toxicity. In some regions (e.g., Saskatchewan, Quebec and Atlantic), Mn concentrations exceeded the health-based guideline value (0.12 mg/L), raising concerns about neurological effects from prolonged exposure, especially among sensitive populations such as children. U concentrations approach or exceed the MAC (0.02 mg/L) in select regions of Saskatchewan and Ontario, reflecting geogenic impact and reinforcing the importance of GW-specific risk assessment.

A case study of two small fly-in communities, Garden Hill and Wasagamack First Nation, in northeast Manitoba detected exceedances of contaminants, including toxic metals (As, Pb, Cr, Zn and Cu), in community soil and surface water samples, where solid waste was inappropriately dumped and incinerated. Burning domestic waste and hazardous waste (e.g., e-waste, plastic materials) near homes in the absence of waste collection or recycling services, sanitary landfills and recycling programs was the major cause of soil and surface water contamination in communities. Governmental negligence and regulatory limitations have led to improper solid waste management and water quality aggravation in Indigenous communities (Zagozewski et al., Reference Zagozewski, Judd-Henrey, Nilson and Bharadwaj2011; Oyegunle and Riddell, Reference Oyegunle and Riddell2016). Lack of funding and constant road blockages have also led to the failure of community initiatives to collect, store, ship or recycle waste (Oyegunle and Thompson, Reference Oyegunle and Thompson2018).

Across the reviewed studies, GW emerges as the primary source of water associated with elevated concentrations of As and Mn, reflecting natural background conditions rather than point-source contamination. In contrast, surface water contamination is more commonly associated with localized anthropogenic inputs, including mine tailings, industrial residues and improper solid waste management, as reported in studies from northern Saskatchewan, the Northwest Territories and Ontario. From a public health perspective, As and Pb represent the most critical HMs, due to their toxicity at low concentrations and well-established links to chronic health outcomes. In contrast, Fe and Mn are detected at higher frequencies and concentrations but primarily affect aesthetic quality, infrastructure performance and user acceptance, rather than posing immediate health risks at typical levels. The recent reclassification of Mn as a neurotoxicant in the GCDWQ demonstrates growing recognition of its potential health relevance, particularly for children, although most reported concentrations remain below this threshold.

Contaminants of emerging concern

Other prevalent contaminants in SRR communities

Water quality in SRR communities is often compromised by the co-occurrence of multiple contaminant classes rather than isolated exceedances of individual parameters, complicating monitoring and treatment. Microbial indicators, metals, nutrients and select organic contaminants often co-occur, reflecting common vulnerabilities in source waters. A study on fuzzy synthetic evaluation of drinking water quality investigated contaminants that caused 126 water advisories in British Columbia in SRR communities and found that in many cases, advisories could not be attributed to a single contaminant. In 61% of cases, the cause of the water advisory was not determined. In 49% of cases, coliforms, Mn, As and Fe led to water quality issues and advisories. Health effects were categorized from undesirable aesthetic effects, including high turbidity and unpleasant taste and odor, to serious carcinogenic risk (Hu et al., Reference Hu, Mian, Abedin, Li, Hewage and Sadiq2022).

While microbial contamination is a well-documented concern, similar co-occurrence patterns are reported in GW-dependent First Nations communities in Ontario. Marshall et al. investigated rapid transport of high levels of nitrate, E. coli, total coliforms, PPCPs and artificial sweeteners in GW of First Nations reserves in Ontario. Using septic systems and lagoons for wastewater collection near-surface and fractured-rock aquifers posed a critical threat to communities, as well as to GW wells at various aquifer depths (Marshall et al., Reference Marshall, Levison, McBean and Parker2019). These findings demonstrate that contaminants traditionally considered separately can share common transport pathways, particularly in hydrogeologically vulnerable settings. GW in SRR contexts appears vulnerable to multi-contaminant intrusion, especially where aquifers are shallow, fractured or hydraulically connected to surface activities. Septic systems, wastewater lagoons, agricultural practices and improper waste disposal facilitate the downward migration of nitrates, fecal bacteria and mobile organic compounds, leading to the co-occurrence of chemical and microbial contaminants in source water.

Another research in 2014 reviewed works done between 1990 and 2013 on GW contamination in North America (USA and Canada). Over 102 private domestic wells were examined in 55 studies. Forty-one samples were collected from 15 locations across 5 provinces in Canada (Goss et al., Reference Goss, Barry and Rudolph1998; Fitzgerald et al., Reference Fitzgerald, Neilson and Kiely2001; Budu-Amoako et al., Reference Budu-Amoako, Greenwood, Dixon, Barkema and McClure2012a; Budu-Amoako et al., Reference Budu-Amoako, Greenwood, Dixon, Barkema and McClure2012b). Total coliforms were detected in seven locations, E. coli in 10 locations, intestinal enterococci in five locations and coliphage in two locations (Hynds et al., Reference Hynds, Thomas and Pintar2014). Few studies have been published on GW contamination by pathogens. Therefore, it is important to investigate present data gaps.

Source water contamination and traditional food chain interactions

Traditional food sources offer cultural and nutritional benefits to Indigenous communities compared to market-based and processed food. However, evidence from several studies showed that traditional foods consumed by First Nations living on reserves, including plant and animal species, may contain elevated levels of HMs, persistent organic pollutants and bioaccumulative chemicals (Savard et al., Reference Savard, Bégin, Parent, Marion and Smirnoff2006). Traditional food sources and drinking water sources are often exposed to shared contamination pathways, resulting in concurrent risks to both food and water security.

Many traditional foods, including freshwater fish, waterfowl and semi-aquatic mammals, are harvested from lakes, rivers and wetlands that also serve as drinking water sources or recharge zones for community water supplies. Consequently, metals and persistent organic pollutants detected in traditional foods, such as Hg, Pb and polychlorinated biphenyls (PCBs), frequently originate from the same upstream drivers affecting drinking water quality, including mining residues, agricultural runoff and wastewater discharges (Sweta and Singh, Reference Sweta and Singh2024). Long-range transported contaminants enter aquatic systems and subsequently bioaccumulate through food webs, impacting both harvested foods and/or inadequately treated drinking water.

Shared exposure pathways have direct implications for social acceptance and risk perception around drinking water and traditional food insecurity. Communities experiencing recurrent water quality deterioration and advisories, or persistent aesthetic concerns, may develop distrust in tap water that extends to traditional foods harvested from local ecosystems (Dewailly, Reference Dewailly2006). The gradual shift from traditional foods to processed and Western foods over the last three decades has been associated with deteriorating health and increasing prevalence of obesity and diabetes among Indigenous populations in the Arctic (Bordeleau et al., Reference Bordeleau, Asselin, Mazerolle and Imbeau2016).

As demonstrated in Table 3, previously conducted works have examined the levels of prevalent contaminants, such as Hg, Cd, Pb, methylmercury, persistent bioaccumulative toxic substances, PCBs and Se in traditional food (fish species, wild food, plants, etc.) of First Nations communities from 10 Canadian provinces and eight Assembly of First Nations (Chan et al., Reference Chan, Singh, Batal, Maruska, Tikhonov, Sadik, Schwartz, Ing and Fediuk2021), Anishinaabe tribal fisheries in Lake Huron, Lake Superior and Lake Michigan (Dellinger et al., Reference Dellinger, Olson, Holub and Ripley2018), the Great Lakes region for the Ojibwa Native Americans (Chiu et al., Reference Chiu, Beaubier, Chiu, Chan and Gerstenberger2004) and Wapekeka and Kasabonika Lake in Northern Ontario (Seabert et al., Reference Seabert, Pal, Pinet, Haman and Robidoux2014). Studies suggest that, in most cases, contaminant concentrations in animal tissue samples collected from nature (e.g., lakes and rivers) were likely higher than those in commercial brands in grocery stores (Dellinger et al., Reference Dellinger, Olson, Holub and Ripley2018).

Table 3.

Studies on traditional food source contamination in indigenous communities in Canada

Moreover, CECs add another layer of concern in the Arctic. MPs, biota (including Arctic char) and PPCPs are being detected in Arctic waters and bioaccumulate in Arctic food webs. Food sources in the Arctic reportedly contained higher levels of PCBs and other tested contaminants. The majority of Arctic contamination results from long-range transport of pollutants from lower latitudes, along with ecological interactions and biochemical processes (Basu et al., Reference Basu, Abass, Dietz, Krümmel, Rautio and Weihe2022).

FNFNES is a nationwide research initiative in Canada responsible for collecting traditional food samples (e.g., wild freshwater and marine fish, land mammals and wild birds) for various contaminant analyses. In November 2019, FNFNES published a comprehensive report on key findings from primary traditional food sources in eight Assembly of First Nations regions comprising 583 First Nations communities and levels of exposure to HMs, PPCPs and PCBs in food and tap water. This decade-long study on traditional foods in different regions of Canada served as a guideline for the aforementioned studies, which compared contaminant concentrations and intake among First Nations (FNFNES, 2019). According to this document, although traditional food sources are generally safe to consume and nutritionally valuable, specific long-lived predatory fish and bird species samples contain high levels of Hg and Pb. FNFNES also demonstrated regional co-occurrence of contaminants in both food and tap water, reinforcing the role of interconnected source water pathways and land-use pressures in shaping cumulative exposure risks.

Analysis of contaminants studied in SRR communities

A comprehensive analysis of the studies and reports investigating different contaminants in SRR communities is shown in Figure 3(a). All publications referenced in this paper on contaminant analysis are listed in Supplementary Table S1, in the supplementary data. Among 36 publications (2001–2025) categorized into three groups, 61% were on HMs and trace minerals, 26% on emerging contaminants and 13% on microbiological contaminants. As (22.6%), Pb (6.5%) and Mn (4.8%) accounted for 33.9% of the publications, reflecting their frequent detection in SRR water supplies. In the CECs category, MPs were tested more than PPCPs and PFAS. Fecal bacteria, including E. coli and total coliforms, were also the most commonly reported pathogenic microorganisms in the literature.

Figure 3.

(a) The most frequently tested contaminants in SRR communities in Canada by a dedicated percentage of studies, and (b) Distribution of studies on three contaminant categories among different provinces.

Protozoa and cysts in the microbial category, along with U, Al, Cd, Hg, Na, Cr, Ag, Co and Ni in the HMs category, accounted for only 1.6% of the documented research, signifying limited attention in the academic literature.

Figure 3(b) demonstrates the distribution of studies among different provinces in three categories: microbial (E coli, total coliform, pathogenic microorganisms, protozoan parasites, cysts and Cryptosporidium), HMs and trace minerals (As, Pb, U, Al, Cu, Cd, Hg, Fe, Mn, Na, Zn, Ag, Co, Ni, Sb and Cr) and CECs (PPCPs, PFAS, MPs). According to several research studies in Ontario and Manitoba, all three categories were examined in SRR communities in both provinces. Ontario, with 143 tested frequent emerging contaminants, had the highest rate of dedicated studies among provinces. The map illustrates that scholars have paid substantial attention to CECs in the source water of SRR communities, as they have been tested and detected across most provinces. Although many BWAs are caused by E. coli and total coliform, the number of publications on microbiological parameters remains limited.

Multi-barrier SWSs and real-world scenarios

Multi-barrier water systems are widely implemented in SWS applications to manage complex and variable water quality risks. By integrating multiple treatment and protection steps, these systems aim to provide redundancy and resilience against both microbial and chemical contaminants. They involve multiple primary and secondary treatment stages, disinfection stages and protective measures to address a wide range of potential contaminants (Reid et al., Reference Reid, Igou, Zhao, Crittenden, Huang, Westerhoff, Rittmann, Drewes and Chen2023).

Filtration followed by disinfection is the most common technology used in multi-barrier systems. In primary treatment, activated carbon or slow sand filtration is used to remove finer particles, microbes, dissolved substances and chemicals. This stage improves the taste and odor of water (Karim et al., Reference Karim, Guha and Beni2020; Zamyadi et al., Reference Zamyadi, Glover, Yasir, Stuetz, Newcombe, Crosbie, Lin and Henderson2021). Although membrane technologies offer advantages, such as removing a wide range of contaminants, they often require significant energy inputs and higher initial costs. Filter clogging, slow treated water flow rate, periodic maintenance (e.g., cleaning, backwashing and media replacement) and limited removal efficiency for certain contaminants are some of the filtration drawbacks (De Souza et al., Reference De Souza, Pizzolatti and Sens2021). Proper pretreatment is essential to prevent membrane fouling and clogging (Rezakazemi et al., Reference Rezakazemi, Khajeh and Mesbah2017; Cevallos-Mendoza et al., Reference Cevallos-Mendoza, Amorim, Rodríguez-Díaz and Montenegro2022).

Common disinfectants in secondary stages, such as chlorine, ultraviolet (UV) irradiation or ozonation, are employed to kill or inactivate harmful microorganisms (Costa Terin et al., Reference Costa Terin, Luíza Souza Freitas, de Melo Nasser Fava and Patricia Sabogal-Paz2021). Disinfection technologies must be accompanied by pretreatment processes to prevent fouling and remove particulates that can shield microorganisms (Brooks et al., Reference Brooks, Tenorio-Moncada, Gohil, Yu, Estrada-Mendez, Bardales and Richardson2018). Chlorination is the most widely used disinfection method. Although chlorination is a reliable and simple method, it can form disinfection by-products (DBPs) when residual chlorine reacts with organic matter in water. Some DBPs are potentially harmful to the human body and are regulated (Mian et al., Reference Mian, Chhipi-Shrestha, Hewage, Rodriguez and Sadiq2020).

In contrast, UV disinfection is a physical treatment that uses UV light to inactivate or destroy harmful microorganisms; however, it is less likely to provide residual protection (Van Nevel et al., Reference Van Nevel, Koetzsch, Proctor, Besmer, Prest, Vrouwenvelder, Knezev, Boon and Hammes2017). Ozonation is also a chemical-free process that breaks down into oxygen after treatment, leaving no residual chemicals in the water. Ozone must be generated on-site due to its short half-life, which requires energy, specialized equipment and proper maintenance. Other disinfection methods, such as boiling and solar disinfection, are common in developing countries. However, they demand additional physical effort, and their effectiveness depends on multiple external factors (e.g., turbidity, heating time, degree and intensity of sun exposure) (McGuigan et al., Reference McGuigan, Conroy, Mosler, du Preez, Ubomba-Jaswa and Fernandez-Ibañez2012).

In practice, multi-barrier SWSs have been employed in SRR and Indigenous communities to improve drinking water quality under diverse and often challenging conditions. Some case studies illustrate how these systems are adapted, modified and operated to address site-specific water quality issues. The modification and optimization of SSFs in two SRR communities in Saskatchewan were comprehensively studied. Table 4 demonstrates the results before and after adjustments to SFFs. Despite some limitations of SSFs, the design of the polyethylene tank-based SSF system encompassing pre- and posttreatment processes such as ozone oxidation, roughing and biological activated carbon filters improved operation, maintenance and efficiency in turbidity, HMs (Fe and Mn), microorganisms and color removal (Gottinger et al., Reference Gottinger, McMartin, Price and Hanson2011).

Table 4.

Reported removal percentage achieved by conventional and modified SFF technology

Note: BAC: biological activated carbon; GAC: granular activated carbon; NTU: Nephelometric Turbidity Units; TCU: true color units; DOC: dissolved organic carbon.

In particular, turbidity removal increased to up to 99.7% with ozone pretreatment and biological activated carbon (BAC), compared to the operational limit of <1 NTU in conventional systems. Similarly, color removal improved from 25–40% to 81–91% with biological roughing filtration. Enhanced removal efficiencies were also observed for Fe and As, reaching up to 95% and 92%, respectively, under modified configurations. Microbial removal was consistently high, with E. coli and total coliform reductions approaching 98–99% when horizontal flow roughing filters were incorporated. However, conventional and modified SSF are not able to reduce hardness and alkalinity. The study also elaborated on details of the designs and modifications to conventional SSFs, such as adding nonwoven synthetic fabrics to the sand bed surface and pebble matrix filter.

In response to growing water insecurity concerns, such as deteriorating source water quality in Newfoundland and Labrador, Water Canada has adopted and installed small-scale potable water dispensing units (PWDUs) known as water kiosks in 32 SRR and Indigenous communities since 2000 (Water Canada, 2022). This work was carried out in collaboration with Memorial University, the University of Guelph, Municipalities of NL, NunatuKavut Community Council and Nunatsiavut Government, with the support of Qalipu First Nation (NL Government, 2014). Such water systems are customized to suit the specific characteristics of source water and the prevailing conditions within communities. They are designed to be centrally located in the community and give residents access to clean drinking water, often from a central facility. Water kiosks incorporate multiple treatment processes, including ozonation, multimedia anthracite and sand filtration, activated carbon filtration, reverse osmosis (RO) and UV disinfection. Water may be collected directly from these units or delivered to residents.

Furthermore, these kiosks can be designed to operate off-grid by incorporating solar panels (Eger et al., Reference Eger, Minnes, Hudson, Vodden, Parewick and Walsh2021). Despite the benefits of PWDUs, some drawbacks are associated with these technologies. Limited supply chains and service interruptions linked to transportation delays were among the logistical issues faced by PWDUs. They can be subject to increasing running costs, such as piping or delivering PWDU water and the need for backup supplies. PWDU water lacks free chlorine residuals, increasing the risk of recontamination upon exposure to pathogens (Wright et al., Reference Wright, Sargeant, Edge, Ford, Farahbakhsh, Shiwak, Flowers and Harper2018).

In addition to conventional systems, several advanced cost-effective technologies have been developed and tested at the pilot scale in SRR communities. One of the initiatives launched in 2009, the RES′EAU WaterNet program, aimed to address existing drinking water challenges and lift water advisories in First Nations communities with populations of up to 500 residents (Cook et al., Reference Cook, Brown, Higashitani and Mohseni2017).

A POE low-pressure UV disinfection system employing three multimedia cartridge filtration units was implemented and tested by the RES′EAU WaterNet program over 1 year in Lytton First Nation in Fraser Canyon, British Columbia. The system served five households (56 residents) with a capacity of 38 LPM. The community was under long-term BWA due to high turbidity in its primary surface water source. The system reduced DOC, turbidity, total coliforms and E. coli to levels compliant with GCDWQ guidelines. The POE project prioritized early community engagement and on-site visits to understand site-specific attributes, present the project to residents and assess homeowner interest. The pilot POE project was managed by the participating First Nation community to support the effective preparation phase. Despite successful public engagement efforts, seasonal isolation compromised system reliability, particularly the long-term feasibility of timely lamp replacement and maintaining a consistent fuel supply for electricity generators (Moore et al., Reference Moore, Pousty, Pras, Gehr, Wong, Ma, Linden, Hofmann, Mamane and Beck2023).

In response to high concentrations of Mn and Fe in GW in Oneida First Nation of the Thames, a three-step treatment system including pre-chlorination, two units of Mn greensand filtration coated with potassium permanganate for Mn and Fe ionization and removal, followed by sodium hypochlorite disinfection was effectively employed (Plummer et al., Reference Plummer, de Grosbois, Armitage and de Loë2013). In the Mississaugas of the Credit First Nation, with surface water (Lake Erie) as the main source, treatment systems include pre-chlorination, clarifying process by coagulation, filtration, flocculation and sedimentation and taste and odor control. High turbidity, pathogens and invasive zebra mussels were the main concerns in the community’s source. If necessary, a bleach or UV system may be added for disinfection, along with a softener to reduce total hardness (Goretsky, Reference Goretsky2021).

Beyond technical design, broader management and operational measures play a critical role in system efficiency. To safeguard the implementation of multi-barrier systems, source water protection, regular monitoring and contingency plans to respond to unexpected events must be accompanied by technology (Mohr et al., Reference Mohr, Dockhorn, Drewes, Karwat, Lackner, Lotz, Nahrstedt, Nocker, Schramm and Zimmermann2020). Multi-barrier systems provide comprehensive protection against a wide range of contaminants, resulting in high-quality drinking water with improved taste and odor. Having multiple treatment stages, if one stage fails in the system, additional barriers in place provide redundancy (Sudhakaran et al., Reference Sudhakaran, Maeng and Amy2013). However, multi-barrier systems can be complex and costly to implement and maintain. Treatment stages require energy and chemical use. They can be resource-intensive because proper operation and continuous monitoring require skilled personnel and resources (Freitas et al., Reference Freitas, Terin and Sabogal-Paz2023).

SWSs’ capacity in priority CECs removal

While multi-barrier SWSs are effective in managing conventional water quality parameters, their capacity to address CECs remains less well characterized in SRR contexts. Given limited data on CEC occurrence and SWS performance, evidence from established prioritization and monitoring programs can be transferred to support CEC removal in SRR-specific constraints. Scholars employ CEC screening tools, including structured chemical prioritization frameworks (e.g., exposure–hazard–risk indicators), multi-criteria ranking methods and network-based approaches such as the NORMAN prioritization scheme (Dulio et al., Reference Dulio, Alygizakis, Ng, Schymanski, Andres, Vorkamp, Hollender, Finckh, Aalizadeh, Ahrens, Bouhoulle, Čirka, Derksen, Deviller, Duffek, Esperanza, Fischer, Fu, Gago-Ferrero and der Ohe2024; Zhao et al., Reference Zhao, Guo, Yang, Liu, Zhang, Luo, Zhang, Wang, Chen and Xu2024). CECs are prioritized based on occurrence, persistence, mobility, ecological and human health risk, exposure and bioaccumulation metrics.

Across international networks and drinking-water risk-ranking studies, several CEC groups consistently emerge as a high priority for targeted monitoring in communities. PFAS, in particular, is persistent and difficult to degrade and is widely recognized as requiring adsorption (e.g., GAC, RO), ion exchange or nanofiltration (NF) and ultrafiltration (UF) for effective removal (Patterson et al., Reference Patterson, Burkhardt, Schupp, Krishnan, Dyment, Merritt, Zintek and Kleinmaier2019). For trace organic contaminants, including PPCPs and certain pesticides, standardized third-party certification (e.g., NSF/ANSI 401 for emerging contaminants) provides a practical evidence pathway for selecting POU/POE systems with verified reduction claims, where communities rely on decentralized treatment. For MPs, the literature indicates that removal is generally size-dependent. Particle-focused technologies (e.g., GAC filtration and membrane filtration) can substantially reduce MPs, supporting their inclusion as a priority for monitoring and barrier evaluation (Na et al., Reference Na, Kim, Kim, Jeong, Lee, Chung and Kim2021).

Priority CECs are not limited to PFAS/PPCPs. Beyond the headline CECs, multiple studies and agency reports highlight additional priority classes relevant to drinking water that are often missed in SRR monitoring programs. The top risk-ranked chemicals include multiple unregulated and under-monitored contaminants across classes in many SWSs, including N-nitrosodimethylamine (NDMA), 1,4-dioxane, chlorate and PFAS compounds (PFOS, PFOA, PFHxS), DBPs and industrial solvents (Rosenblum et al., Reference Rosenblum, Liethen and Miller-Robbie2024). EPA Contaminant Candidate List 5 includes dozens of chemicals and explicitly flags PFAS, cyanotoxins, organophosphate flame retardants (OPFRs) (e.g., TCEP, TCPP/TCIPP) and DBPs as candidate groups for future regulatory determination (EPA, 2022). In the European Commission, the Drinking Water Directive watch list requires monitoring of endocrine disruptors 17-β-estradiol and nonylphenol in drinking water across the supply chain, providing a policy-relevant signal that endocrine-active compounds remain priority targets (EC, 2022).

Table 5 illustrates the capability of POU and POE technologies in SRR communities to remove priority classes of CECs. Findings from drinking water risk-ranking and treatment evaluation studies consistently reported that conventional SWSs, such as basic filtration, chlorination and/or UV disinfection, are not designed to remove most dissolved CECs, including PFAS, nitrosamines, mobile industrial organics (e.g., 1,4-dioxane) and endocrine-disrupting compounds (Rosenblum et al., Reference Rosenblum, Liethen and Miller-Robbie2024). Adsorption-based processes and ion exchange provide broader but capacity-limited removal for selected organic CECs and PFAS, while membrane processes (RO/NF) offer the most robust multi-contaminant barrier at a small scale (PHO, 2025). Some multi-barrier systems and PWDUs incorporate GAC, ozonation or RO, but these technologies are rarely deployed with explicit CEC mitigation objectives or accompanied by monitoring frameworks capable of verifying CEC reduction under field conditions.

Table 5.

SWSs’ capacity in priority CECs removal

Note: ✓✓ = strong evidence/generally effective ✓ = effective but context-dependent Δ = variable/limited/compound-specific ✗ = generally ineffective or not designed for this purpose.

In contrast, oxidation-based processes, including UV-based advanced oxidation process (AOP) and electrochemical oxidation (EO), are effective for highly mobile and poorly adsorbing contaminants such as NDMA and 1,4-dioxane, but remain limited by energy demand, by-product formation and feasibility for decentralized deployment (Harish and Jegatheesan, Reference Harish and Jegatheesan2025). Particle-associated CECs, including cyanobacterial cells and MPs, are best addressed with physical separation processes such as coagulation and filtration, underscoring the importance of distinguishing dissolved versus particulate CECs in prioritization criteria.

CECs’ prioritization should integrate decision-oriented approaches focused on SWS treatment capabilities. In terms of prioritization, multi-criteria frameworks should be extended by weighting POU/POE feasibility and effectiveness for these priority groups with regard to SWS constraints. Priority metrics should also account for mixture effects, particularly for PFAS co-occurring with other organic micropollutants, by using mixture indicators (e.g., ΣPFAS, bioassay-based endocrine activity proxies) and evaluating them against the strengths and limitations of available POU/POE technologies. On the treatment side, SWS metrics can be studied through comparative evaluations of field-relevant performance data (e.g., media life, breakthrough behavior, high natural organic matter (NOM), varied turbidity, intermittent use). For oxidation-based small-scale systems (e.g., UV/ozonation), future studies should emphasize transformation products, by-product formation and operational feasibility. Finally, while third-party certification frameworks (e.g., NSF/ANSI 401, 53 and 58) provide a practical starting point for device selection, their claims should be validated under SRR-specific conditions, including extreme cold climate, operation and maintenance constraints, variable water chemistry and realistic usage patterns.

SWS performance in extreme cold climates

Beyond system configuration and removal capabilities, SWSs must function reliably under harsh environmental conditions, particularly in cold climates. The evaluation of SWSs under extreme cold conditions is important in Canada, where many SRR and Indigenous regions experience prolonged sub-zero temperatures. Extreme cold can alter source-water quality, slow treatment kinetics and challenge operation and maintenance, potentially reducing treatment reliability. Cold climate affects SWSs through two coupled pathways: temperature-dependent process constraints (e.g., slower reaction kinetics, higher viscosity and altered contaminant speciation/particle dynamics) and technical constraints (distribution, storage, intermittent use, limited operator access, etc.).

Arctic Canadian field studies of surface water sources in small fly-in communities supplied by SSF and UV disinfection, in addition to chlorination, show that even when source water quality is acceptable, water quality can deteriorate between treatment and tap because of challenges in maintaining disinfectant residuals during treatment and distribution. More studies on decentralized drinking water systems in Arctic regions indicate lower microbial activity, turbidity and NOM levels in both source and stored water, but higher chlorine residual non-compliance (Gora et al., Reference Gora, Soucie, McCormick, Ontiveros, L’Hérault, Gavin, Trueman, Campbell, Stoddart and Gagnon2020; Duncan et al., Reference Duncan, Ibraheem, Tam, Egotak, Maniyogina, Klengenberg and Gora2025). In several Nunavut communities, plumbing and storage conditions contributed to Fe, Mn and occasional Pb exceedances, while increased corrosion in distribution pipes during colder periods led to aesthetic water quality deterioration (Daley et al., Reference Daley, Castleden, Jamieson, Furgal and Ell2015).

Many researchers explored the efficiency of SSF in SRR communities in Canada. As part of the Ontario Safe Drinking Water Act, multistage sand filtration followed by pre-ozonation and post-GAC technology was implemented in a small northern community. In cold temperatures as low as 2°C, microorganisms’ activities and the need for biological treatment decrease (Shi et al., Reference Shi, Ma and Zhang2022). Cold water temperature within the microfiltration range can impact filtration performance. In cold temperatures, water velocity and hydraulic flow rate decline, resulting in a more resistant, less flexible and less permeable environment (Partinoudi et al., Reference Partinoudi, Collins, Dwyer and Martin-Doole2007; France et al., Reference France, Bot, Kelly, Crowley and O’Mahony2021). This leads to filtration clogging and fouling. Therefore, frequent filter cleaning and maintenance are required. To address this issue, the filtration system was adjusted by adding two different depths of roughing sand filters. The roughing filter played an important role in preventing solid loading and particle accumulation on the slow sand filter. It also significantly enhanced the removal of turbidity, total coliform and cryptosporidium parvum oocysts when the system had not yet attained biological maturity (Cleary, Reference Cleary2005).

For membrane-based SWSs (UF/NF/RO), cold water increases viscosity and can intensify fouling and cleaning challenges. Reviews and experimental studies report flux/permeability losses and increased resistance under cold temperatures, implying higher energy demand or reduced production capacity during winter conditions unless systems are re-optimized (flux reduction, cleaning adjustments, etc.) (Xu et al., Reference Xu, Gao, Liao, Bai, Qiao and Turek2023).

Extreme cold conditions increase the vulnerability of treatment processes and overall system performance, highlighting a critical evidence gap in supporting climate-resilient SWS design and management. In cold-region and Arctic communities, prolonged sub-zero temperatures pose operational and maintenance challenges, including maintaining adequate disinfectant residuals, preventing freezing and equipment damage, accessing treatment units for routine servicing and ensuring timely cartridge or media replacement, often under limited technical capacity and supply chain disruption.

SWS removal performance for prevalent contaminants in SRR communities

Analysis of contaminant occurrence patterns in SRR and Indigenous communities provides a foundation for evaluating whether SWSs are appropriately aligned with dominant water quality risks in practice. Real-world SWS implementation in SRR communities indicates that most systems are designed primarily to address microbial risks and water quality parameters, rather than dissolved inorganics or priority CECs. Multi-barrier configurations commonly rely on combinations of filtration (e.g., SSF, multimedia, cartridge filtration) followed by chlorination and/or UV disinfection. These systems have demonstrated strong effectiveness in achieving microbial compliance at the point of treatment, particularly for E. coli and total coliforms, and in reducing turbidity to levels that support effective disinfection. However, multiple field studies summarized in this review demonstrate that water quality can deteriorate between treatment and the point of consumption due to recontamination during storage, transport and distribution, compounded by inconsistent disinfectant residuals. Technologies such as UV disinfection, which do not provide residual disinfection, make the treatment-to-tap journey vulnerable to recontamination. This limitation is particularly relevant in SRR contexts, where decentralized storage and transportation are common.

Across the reviewed studies, Fe and Mn repeatedly occurred at elevated concentrations in GW, largely due to natural background conditions. While historically treated as aesthetic concerns, elevated Fe and Mn levels impose significant operational burdens on SWSs through filter clogging, membrane fouling, staining and taste and odor complaints. The formation of Fe and Mn oxide deposits also leads to fouling in pipes, storage tanks and treatment infrastructure. Consequently, several employed SWSs prioritize Fe/Mn oxidation and removal using processes such as pre-chlorination or permanganate oxidation followed by greensand or media filtration. These systems demonstrate effective performance in reducing prevalent water quality issues identified in Section 3. In practice, the inclusion of additional treatment barriers in these systems primarily serves to stabilize operation and maintain consumer acceptability, rather than explicitly targeting a broader range of chemical hazards.

In contrast, As, despite being one of the most frequently reported contaminants in SRR GW supplies, appears less consistently addressed in the real-world SWS case studies highlighted in this review. While As mitigation typically requires adsorption-based media, ion exchange or membrane processes such as RO or NF, fewer implemented systems are framed or evaluated around As removal objectives. This suggests a potential misalignment between contaminant prevalence and treatment focus, which may reflect feasibility constraints, cost considerations and monitoring limitations. The gap is notable given that conventional filtration and disinfection processes are not designed to address dissolved As species. As in the community, GW is commonly of natural origin and associated with local geology and aquifer geochemical conditions. When elevated As concentrations are identified, public health guidance emphasizes the use of an alternate source water as a primary mitigation strategy, particularly where treatment options are limited or operationally challenging (FH, 2013). In communities with decentralized systems, a management response is source water substitution, such as switching to an alternative or deeper aquifer with lower As levels by installing a new well (George et al., Reference George, Inauen, Perin, Tighe, Hasan and Zheng2017).

Compared with decentralized, small-scale water treatment approaches used in other SRR and rural contexts globally, Canadian SRR deployments demonstrate a strong emphasis on multi-barrier microbial protection. International frameworks for household and decentralized water treatment similarly prioritize filtration and disinfection as the foundation of safe water provision. Overall, the synthesis of contaminant occurrence and real-world SWS implementation indicates that current systems are generally well aligned with the most immediate and operationally visible water quality challenges in SRR communities, particularly microbial contamination, turbidity and Fe/Mn-related issues. However, alignment weakens for contaminants that are chronic, less perceptible or analytically complex, including As and priority CECs. Addressing this gap will require closer integration of contaminant prioritization with technology selection, expanded monitoring beyond conventional compliance parameters and evaluation of SWS performance under the realistic operational, climatic and infrastructural constraints.