A Critical Review of Domestic Wastewater Pollutants: Exposure Pathways and Treatment Technologies

Abstract

Domestic wastewater is a chemically complex and highly variable mixture of pollutants generated by everyday household activities, yet its contribution to environmental contamination is still frequently underestimated and only 56% of wastewater worldwide is being treated. This review provides a structured and quantitative assessment of major domestic wastewater pollutant groups, their principal exposure pathways, and current and emerging treatment technologies. Beyond a conventional narrative synthesis, the review derives per capita annual emission estimates from published data and uses these to compare pollutant groups by mass flow and environmental relevance. The analysis shows that high-volume household inputs, particularly sodium chloride from domestic water softening, toilet paper, personal-care products, detergents, and cleaning agents, can contribute substantially to overall pollutant loads, whereas lower-mass contaminants such as pharmaceuticals, antibiotics, PFAS, heavy metals, and microplastics remain critical because of their persistence, biological activity, and incomplete removal during treatment. The review further highlights that conventional wastewater treatment systems are often poorly equipped to remove many of these emerging contaminants effectively, especially under decentralised or only partially advanced treatment conditions. Advanced and hybrid technologies, including membrane bioreactors, nanofiltration, reverse osmosis, adsorption, photocatalysis, and electrochemical processes, offer clear potential, but their broader implementation remains constrained by cost, energy demand, fouling, and concentrate management. Overall, the added value of this review lies in linking mass-based pollutant prioritisation with treatment performance, thereby providing a more systematic basis for identifying dominant household emission pathways and for guiding targeted mitigation and technology selection in future wastewater management.

1. Introduction

1.1. Domestic Wastewater Pollutants

Domestic wastewater is a chemically heterogeneous mixture of organic and inorganic substances generated by routine household activities, including water softening, laundry, cleaning, cooking, and the use of medications and personal hygiene products. In contrast to industrial wastewater, which typically has a defined and comparatively stable composition, domestic wastewater exhibits substantial variability and unpredictability on both short (hourly) and longer (seasonal) timescales [1]. These fluctuations are largely driven by human behaviour, consumption patterns, and climatic conditions.

Although most contaminants occur at low concentrations relative to industrial effluents, many display high biological activity and/or a propensity to bioaccumulate, resulting in long-term ecological relevance despite trace levels [2,3]. The following sections summarise the major pollutant classes in domestic wastewater, their environmental impacts, and current technological approaches for their removal from water systems, before concluding with a detailed discussion of specific treatment mechanisms and process efficiencies.

In addition to identifying key pollutant groups, this review also examines currently available treatment technologies and mitigation strategies aimed at reducing their environmental impact.

1.2. Water Softening

Many households, particularly in hard-water regions, use water softeners to remove calcium and magnesium ions responsible for hardness and scale formation. The most common approach is ion exchange, in which Ca²⁺ and Mg²⁺ cations are replaced by sodium ions (Na⁺). While effective for preventing scale formation, this process markedly increases sodium and chloride concentrations in wastewater, because brine generated during regeneration of ion-exchange resins with NaCl is typically discharged to sewers or septic systems, thereby increasing overall salinity [4].

Elevated chloride concentrations impair microbial communities in activated sludge, reducing the efficiency of biological nitrogen removal and organic carbon degradation [5]. Such effects have been reported in both municipal and decentralised (e.g., domestic) wastewater treatment systems. Furthermore, excessive salinity contributes to soil degradation, biodiversity loss, and toxicity to aquatic plants.

To mitigate these effects, several alternative softening technologies have been proposed, including template-assisted crystallisation, membrane nanofiltration, ion exchange using sodium-free resins, and electrode ionisation. Despite technical feasibility, broader deployment remains constrained by cost, which may require targeted policy interventions and regulatory incentives to overcome.

1.3. Cleaning and Detergent Agents

Detergents represent one of the most prevalent pollutant groups in domestic wastewater, reflecting their widespread use in personal hygiene, dishwashing, laundry, and household cleaning. Typical formulations contain complex mixtures of surfactants (anionic, cationic, non-ionic, and amphoteric), builders (including phosphates), bleaches, enzymes, dyes, preservatives, and fragrances [6,7]. Many constituents are only partially biodegradable, and, in some cases, degradation intermediates can be more toxic than the parent compounds [8].

In wastewater, detergents contribute to foaming and substantially increase chemical oxygen demand (COD) and biochemical oxygen demand (BOD), thereby placing additional load on treatment processes. In receiving waters, surfactants can disrupt the membranes, inhibit algal photosynthesis, and act as endocrine disruptors [9]. Phosphates and phosphonates are of particular concern because they can accelerate eutrophication in surface waters. Recent studies suggest that aminopolyphosphonates (used in detergents to prevent limescale) may be converted to glyphosate during biological treatment in wastewater treatment plants [10]. Although household sources may contribute to environmental glyphosate burdens, agricultural use remains the dominant emission pathway.

To address these concerns, plant-based biosurfactants and “greener” detergents formulations are receiving increasing attention. However, the widespread use of disinfectants during and after the COVID-19 pandemic has created additional challenges. These chemically stable, highly antimicrobial compounds can inhibit biological treatment, thereby reducing the performance of conventional wastewater treatment plants.

1.4. Antibiotics

Antibiotics enter domestic wastewater primarily via excretion in urine and faeces and through improper disposal of unused or expired medicines. Although typically present at trace concentrations, antibiotics are biologically active and may persist in aquatic systems [11]. Commonly detected classes include β-lactams, macrolides, fluoroquinolones, and tetracyclines [12]. Because many antibiotics degrade slowly, they can persist for extended periods and accumulate in sediments and biota.

The presence of antibiotics in wastewater promotes the emergence and proliferation of antibiotic-resistant bacteria (ARB) and facilitates horizontal transfer of antibiotic resistance genes (ARGs), particularly within biofilms and sewage sludge [13]. This is a global environmental and public health challenge and has been recognised by the World Health Organization as a major threat to modern medicine.

Conventional treatment processes, including activated sludge, denitrification, and membrane filtration, often achieve insufficient antibiotic removal. Consequently, advanced oxidation and hybrid approaches are being developed and evaluated, including photocatalysis, ozonation, adsorption on porous sorbents (e.g., activated carbon, graphene, and metal-organic frameworks [MOFs]), and a range of electrochemical methods. Alongside technical measures, public education on prudent antibiotic use and appropriate disposal remains essential for a comprehensive mitigation strategy.

1.5. Microplastics in Washing Process

Microplastics have emerged as persistent and difficult-to-control pollutants in aquatic environments [14]. In domestic contexts, a major source is the laundering of synthetic textiles, such as polyester, acrylic, and nylon [15]. Mechanical abrasion and fibre fragmentation during washing and drying release microscopic plastic particles. Estimates indicate that a single wash can release tens of thousands to hundreds of thousands of microplastic fibres, depending on fabric type, garment age, and washing conditions [16].

Conventional wastewater treatment plants do not completely remove microplastics. A fraction is captured in sewage sludge, but sludge is often applied to agricultural soils as fertiliser, creating a secondary pathway of environmental contamination. Microplastics also provide sorption surfaces for heavy metals and organic contaminants, including pesticides and pharmaceuticals [17]. While sorption may temporarily immobilise these compounds, it can also facilitate their transport and redistribution across aquatic and terrestrial environments.

Source-control measures include the integration of fibre-capturing filters in household washing machines. However, these devices can themselves become sources of microplastics if poorly maintained, damaged, or improperly disposed of. Accordingly, current research is exploring improved removal and degradation strategies, including physical separation, coagulation, membrane filtration, photocatalysis, and biological degradation, as well as hybrid process combinations.

1.6. Other Pollutants

Additional contaminant groups commonly detected in domestic wastewater include personal care and pharmaceutical products (PPCPs), such as parabens, preservatives, hormones, ultraviolet (UV) filters, and antioxidants. Many PPCPs exhibit endocrine-disrupting activity, and their metabolites may be persistent and toxic [18].

Heavy metals constitute another important class of concern. They enter domestic wastewater through corrosion or leaching from plumbing materials (e.g., Cu, Zn, and Pb) and via consumer products, including hair dyes, make-up, and skin-care formulations [19,20]. In recent years, substantial attention has also focused on perfluoroalkyl and polyfluoroalkyl substances (PFAS), widely used in fabric protectants, non-stick coatings, and certain cleaning products. Thompson et al. identified toilet paper as a major PFAS source entering wastewater treatment plants [21]. PFAS are of particular concern because of their exceptional persistence, resistance to degradation, and high bioaccumulation potential.

Household fats, oils, and grease (FOG) can also create operational problems in sewer networks and treatment plants. These materials may solidify on cooling, causing blockages, disrupting biological processes, and leading to costly maintenance and infrastructure damage [22].

The coexistence of chemically diverse pollutants creates complex interactions that can yield additive or synergistic toxicity, complicating the prediction of overall environmental impacts from controlled laboratory tests. This multicomponent, time-varying composition is a key reason why domestic wastewater remains challenging to treat effectively using conventional technologies alone.

1.7. Summary and Major Challenges

Domestic wastewater contaminants combine chemical diversity with environmental persistence and biological activity. Even when present at low concentrations, their continuous release and long residence times can lead to gradual accumulation and, ultimately, potentially irreversible changes in ecosystems. Many municipal wastewater treatment plants were designed decades ago for bulk parameters (e.g., suspended solids, BOD, and nutrients) and not for the broad range of emerging contaminants now detected in domestic wastewater.

Key challenges include limited regulatory standards for many newly recognised contaminants and an incomplete understanding of the cumulative toxicity of multicomponent mixtures. Addressing these gaps will require both improved monitoring and the deployment of advanced treatment processes, including adsorption, photocatalysis, ozonation, electrochemical oxidation, and membrane-based separations.

Ultimately, effective mitigation will require an integrated approach that considers the full contaminant life cycle—from generation and release through transport, transformation, and final environmental fate—to reduce long-term risks to human health and aquatic ecosystems.

Although this review aims to provide a global perspective on domestic wastewater pollutants, several quantitative estimates are derived from well-documented datasets from Germany and the EU, largely due to the ready availability of high quality data for these regions. We acknowledge that this may affect the applicability/transferability of our findings to other geographic and socio-economic contexts. Nevertheless, we employ these data as representative benchmarks to estimate orders of magnitude and relative contributions, rather than as universally applicable values.

The objective of this review is to provide a structured and quantitative assessment of domestic wastewater pollutants, their main exposure pathways, and available treatment technologies. The scope includes key pollutant groups originating from household activities, such as detergents, pharmaceuticals, microplastics, and salinity-related compounds, with particular emphasis on their mass flows and environmental relevance.

In contrast to conventional narrative reviews, this review introduces a quantitative perspective by deriving per-capita emission estimates from literature data and ranking pollutant groups according to their annual mass contribution. This approach provides a basis for systematic comparison of the relative importance of different pollutant sources. Furthermore, the quantitative analysis is integrated with the assessment of treatment technologies, thereby linking specific pollutant groups to relevant mitigation approaches.

2. Exposure in the Environment

2.1. Exposure Pathways On-Site

Private households, including single-family and multi-family homes, generate wastewater with a complex and variable pollutant mixture. Pollutants enter domestic wastewater via household drains (e.g., kitchen and bathroom), toilet flushing, the use of washing machines and dishwashers, and the operation of water softeners. Pollutant-containing wastewater may also arise from cleaning indoor spaces and outdoor areas, car washing, and gardening activities. Because ecotoxicological effects can vary substantially among substances and remain an active area of research, comparing input quantities is of particular value.

Table 1. Ranking of environmentally relevant substance groups/compounds by mass per capita per year and principal entry pathway. Quantities refer to substances/products used (not to individual constituents).

Rank	Entry Pathway	Substances/Products	Key Environmentally Relevant Constituents	Amount [kg Person−1 Year−1]
1	Water softening	Salt used in domestic softeners	Sodium chloride	17.57–46.86
2	Toilet flushing	Toilet paper	PFAS	3.4–26
3	Personal care products	Handwash and shower gel, shampoo, soap, toothpaste, sunscreen, cosmetics	Surfactants, parabens, nanoparticles, microplastics, etc.	11.70–21.46
4	Laundry	Detergents and laundry additives	Surfactants, builders	1.99–16.10
5	Household cleaning, car washing, gardening	Cleaning agents, herbicides and pesticides	Surfactants, herbicides and pesticides	2.67–12.91
6	Dishwashing	Dishwashing and cleaning products	Surfactants, builders, microplastics	9.30–9.89
7	Toilet flushing, kitchen sink, dishwasher	Cooking oils and fats	Cooking oils and fats	0.800
8	Laundry	Microplastics released by textile abrasion	Microplastics, dyes/coatings associated with the plastic particles	0.212
9	Improper disposal, laundry of PPE, leaching from pipework/roof drainage, etc.	Heavy metals	Heavy metals	0.162
10	Urine/faeces via toilet flushing; possibly disposal via drains	Pharmaceutical residues	Antibiotics, analgesics, antiepileptics, etc.	0.00075

compares selected substances and product groups, and their environmentally relevant components, in terms of per-capita annual quantities and principal entry pathways. The degradation behaviour and fate of these material flows within the wastewater treatment plant (WWTP) are discussed in detail in Section 2.2. This quantitative comparison, together with the discussion in the present section, is intended to highlight environmentally relevant consumption and disposal practices. Section 2.3 addresses the socio-ecological aspects of household wastewater disposal.

Water softeners are widely used in households and industry. Both centralised and decentralised systems are deployed, but decentralised household units are especially common in hard-water regions and, in many settings, account for most installations [23]. Their contribution to chloride loading can be substantial. For example, in Minnesota (USA), 65% of the chloride load entering wastewater treatment plants has been attributed to commercial and residential softening systems, with private households responsible for 49% of total chloride inputs [24]. Centralised softening also occurs: in the Netherlands, approximately 1200 Mm³ of drinking water is produced annually, and about half is softened to 1.4–1.5 mmol/L [25].

In rural areas, where sewer connections are not always available, wastewater is often treated using on-site systems. In the USA, up to 20 million households and commercial premises rely on septic systems for wastewater treatment [26]. In rural regions of Greece, up to 14% of households use on-site wastewater treatment [27]. Studies of septic tanks frequently report elevated sodium chloride inputs, which have been linked to the use of water softeners. Taken together, these observations indicate that household water softening contributes materially to chloride inputs to wastewater systems, and therefore represents a non-negligible driver of salinity-related environmental pressures.

Figure 1 shows a schematic of a typical ion-exchange softener. Because Ca²⁺ and Mg²⁺ are exchanged for Na⁺, operation increases sodium concentrations in the treated drinking water; typical sodium concentrations are around 100 mg/L [28]. Chloride concentrations in the softened water itself are not increased. For context, the regulatory limits for sodium and chloride in drinking water in Germany are 200 mg/L and 250 mg/L, respectively [29]. These thresholds are generally not exceeded, implying that softened water does not usually represent an immediate health concern. Nevertheless, the broader nutritional relevance of sodium contributed via drinking water remains the subject of ongoing debate and is not considered further here.

During regeneration of the ion-exchange resin, a brine-rich wastewater stream is produced (Figure 1; outlet 2). Reported sodium chloride concentrations typically range from 3 to 45 g/L [30,31]. Chloride accounts for approximately 60% of the mass, corresponding to about 1.8–27 g/L. However, the total volume of regeneration wastewater generated by typical household ion-exchange units is rarely reported because regeneration frequency and water use vary substantially across devices and households. Accordingly, in this review, annual NaCl loading is estimated using literature-reported consumption values and an independent calculation based on normative specifications.

One US-based estimate assumes a salt consumption of 1.8 kg NaCl per day for a water-softening unit. Reported annual consumption values of ~660 kg for a three- to four-person household correspond to 165–220 kg NaCl per person per year [32]. This value appears high and may overestimate typical use, because salt demand depends on influent hardness, set-point hardness, and user behaviour. As a complementary estimate, we therefore derive a plausible range using German normative parameters.

In Germany, manufacturers of salt-based ion-exchange softeners commonly refer to DIN EN 14743 [33]. The standard specifies that, per kilogram of salt used during regeneration, a minimum exchange capacity of 4 mol must be achieved. In practical terms, this corresponds to a salt use of up to 44.575 mg per litre to reduce water hardness by 1 °dH (the standard unit used here for hardness). We adopt this value as an efficiency benchmark.

For per-capita water use, we apply figures from the German Environment Agency (Umweltbundesamt): 120 L per person per day, equivalent to 43,800 L per person per year [34]. Water softeners are commonly installed from a total hardness of ~14 °dH. A typical setting reduces hardness to ~5 °dH, corresponding to a reduction of 9 °dH; we use this as a minimum softening scenario. To define an upper-bound scenario, we consider an initial hardness of 30 °dH (upper end of the “hard” category in a commonly used European classification) and reduction to 6 °dH, yielding a maximum reduction of 24 °dH [35]. Using these assumptions, annual salt consumption per person can be approximated as:

Minimum salt consumption per person per year:

44.575 mg dH⁻¹ L⁻¹ × 43,800 L × 9 °dH = 17.57 kg

Maximum salt consumption per person per year:

44.575 mg dH⁻¹ L⁻¹ × 43,800 L × 24 °dH = 46.86 kg

These estimates are based on normative specifications for ion-exchange water softeners set out in DIN EN 14743, which defines the relationship between salt consumption and hardness reduction.

On the basis of these parameters, a plausible per-capita salt consumption range is 17.57–46.86 kg per year. This range is consistent with empirical estimates reported by Overbo et al. [24], who reported a total annual salt consumption in Minnesota of 11.34 kg per household per month attributable to domestic water softening; assuming and average Minnesota household size of 2 to 3 persons and multiplying by 12 (months) yields a salt consumption of 45.36–68.04 kg per person per year.

Beyond household softening, the use of de-icing salt is a major contributor to chloride loading in many regions. Road salt is applied during winter maintenance to melt snow and ice. EU regulatory and practical use differs across countries: in some, private application is strongly restricted (e.g., Germany, Finland, Slovakia), whereas in others, it is permitted for both public and private use. Pieper et al. estimated a de-icing salt consumption of 62.1 kg per person per year, with the majority of the load associated with public application (e.g., roads and pavements) rather than private use [35]. This additional salt burden compounds environmental impacts beyond those driven by water softening. Moreover, de-icing salt accelerates corrosion of both public and private drinking-water infrastructure [36]. Because municipal authorities are responsible for most de-icing salt inputs, these quantities cannot be directly assigned to household activities; nevertheless, they align with the assessment by Overbo et al., who likewise identify de-icing salt as a significant source of salinity-related pollution [24].

Overall, by mass, salt discharge associated with domestic water softening ranks first among household-related pollutant inputs (Table 1).

Having established the dominant contribution of salt-based water softening, the discussion now turns to other major household-derived inputs that also enter wastewater in comparatively large quantities. One such item is toilet paper consumption which has a substantial environmental footprint, including contributions to deforestation, as well as significant energy, water, and material demands during manufacture [37]. Reported per-capita consumption varies with consumer behaviour and country, ranging from 3.4 (Brazil) to 12.7 (US) kg person⁻¹ year⁻¹ [38], while Thompson et al. reported values as high as 26 kg person⁻¹ year⁻¹ in the US and Canada [21]. More sustainable alternatives include recycled toilet paper and bamboo-based products that avoid virgin wood pulp. However, recent evidence indicates that toilet paper various types can contain contaminants such as per- and polyfluoroalkyl substances (PFAS), suggesting that toilet paper can constitute a relevant PFAS source to wastewater systems [21,39]. Estimated per-capita inputs of specific PFAS, such as 6:2 fluorotelomer phosphate diester (6:2 diPAP), via toilet paper use range from 6.4 to 80 μg person⁻¹ year⁻¹ [21]. Life-cycle assessment (LCA) studies further suggest that more transformative alternatives, such as multifunctional toilets with integrated bidet functions, can reduce overall environmental burdens [40]. These LCA-based perspectives, discussed in more detail in later sections, underscore the value of mass-based accounting of household emission streams (Table 1): high-volume inputs often offer meaningful reduction potential through sustainable consumption choices and/or the targeted deployment of specific treatment technologies.

Other major contributors to domestic wastewater loading are personal-care and protection products, such as shower gels, shampoos, creams, cosmetics, soaps, and toothpaste. Bathing and showering are typically considered the largest components of household water use, accounting for approximately 35–40% [41]. Accordingly, products used for personal care and protection contribute substantial input quantities (Table 1) and can contain constituents that affect both wastewater treatment performance and receiving environments. Table 2 provides an overview of representative product categories and their estimated emissions to wastewater. Product types are intentionally grouped as a highly disaggregated inventory of individual products is beyond the scope of this review. Summing the minimum and maximum values in Table 2 yields an estimated consumption range of 11.70–21.46 kg person⁻¹ year⁻¹.

The estimates in Table 2 were created using the following assumptions and approximations. We use [42,43] as a benchmark, in which quantities entering wastewater are reported in L person⁻¹ year⁻¹. Because most products have densities in the range of ~1.0–1.3 kg L⁻¹, we apply the approximation 1 mL ≈ 1 g. The use of shower gel, shampoo, conditioner, and soap is assumed to occur either daily (upper-bound scenario) or on every second day (conservative scenario), such that per-use mass is multiplied by 365 or 182.5, respectively. Make-up consumption values (in per-use units) from [44] were summed across product types. Because make-up use is predominantly reported for women, and approximately 41% of women report frequent use [44,45], we divided the summed make-up consumption by two (to represent the female population share) and multiplied by 0.41; make-up removers are not included. For moisturising products, only night cream and moisturiser from [44] are considered as representative examples, and again two scenarios (daily use versus application every other day) are considered. As annual comparisons (including periods of active and inactive use) are important for this review, the sunscreen consumption reported by Biesterbos et al. and Gomez-Berrada et al. were considered [45,46]. For DEET (N,N-diethyl-m-toluamide), Bremer et al. assumed 15 applications per year at 6 g per application [47], and these values are applied on an annual basis. For toothpaste consumption from [44], two applications per day are assumed.

It should be noted that a proportion of applied personal-care and protection products is removed by wiping (e.g., with tissues), and, for sunscreens in particular, some fraction may enter aquatic systems directly during use. Nevertheless, we conservatively assume that most of the product mass ultimately reaches wastewater.

Table 2 also summarises representative environmentally relevant constituents in personal-care and protection products. Phthalates, parabens, bisphenols, and antimicrobial agents such as triclosan are well-recognised endocrine-active substances used in cosmetics and personal-care formulations [48]. Cyclosiloxanes, glycol ethers, and ethanolamines can be toxic to aquatic organisms. Fragrances have also been implicated as potential endocrine disruptors and may cause adverse health effects in consumers during use [49,50]. Nanoparticles are employed as UV filters (e.g., zinc oxide or titanium dioxide), as encapsulated active ingredients (e.g., nanocapsules containing retinol), or as fillers (e.g., silica nanoparticles), and such materials can adversely affect aquatic microorganisms [51]. Finally, microplastics are still present in some personal-care products, including body lotions, soaps, and exfoliants [52,53]. Although restrictions and bans have been introduced in some jurisdictions, regulatory implementation remains incomplete, and microplastics therefore remain in certain products.

Table 2. Overview of personal-care and protection products, estimated consumption, and representative environmentally relevant constituents.

Personal Care Products	Estimated Consumption (Kg Person−1 Year−1)	Representative Environmentally Relevant Constituents [52]	Ref.
Hand cleansing gel	4.63	Surfactants, microplastics	[42]
Shower gel	1.06–2.12	Surfactants, microplastics	[43]
	5.11		[42]
Shampoo	2.3	Surfactants, bisphenols, cyclosiloxanes, ethanolamines	[54]
	1.9–3.8		[44]
	4.35		[42]
Hair conditioner	1.82–3.65	Bisphenols	[44]
	3.35		[42]
Soap	0.9	Antimicrobial agents, fragrances, microplastics	[54]
	0.6–1.2		[44]
	0.21		[42]
Cosmetics and deodorant	0.084	Parabens, phthalates, UV filters, microplastics	[44,45]
	0.44 (Deodorant)		[42]
Moisturising cream	0.39–0.78	Parabens, phthalates, UV filters	[44]
	0.42		[42]
Protective creams (sun, insects)	0.146	UV filters, glycol ethers, nanoparticles, insect repellents (e.g., N,N-diethyl-m-toluamide (DEET))	[45,46]
	0.090		[47]
Toothpaste	0.93	Antimicrobial agents, nanoparticles, microplastics	[44]
	1.00		[42]

Table 2. Overview of personal-care and protection products, estimated consumption, and representative environmentally relevant constituents.

Personal Care Products	Estimated Consumption (Kg Person−1 Year−1)	Representative Environmentally Relevant Constituents [52]	Ref.
Hand cleansing gel	4.63	Surfactants, microplastics	[42]
Shower gel	1.06–2.12	Surfactants, microplastics	[43]
	5.11		[42]
Shampoo	2.3	Surfactants, bisphenols, cyclosiloxanes, ethanolamines	[54]
	1.9–3.8		[44]
	4.35		[42]
Hair conditioner	1.82–3.65	Bisphenols	[44]
	3.35		[42]
Soap	0.9	Antimicrobial agents, fragrances, microplastics	[54]
	0.6–1.2		[44]
	0.21		[42]
Cosmetics and deodorant	0.084	Parabens, phthalates, UV filters, microplastics	[44,45]
	0.44 (Deodorant)		[42]
Moisturising cream	0.39–0.78	Parabens, phthalates, UV filters	[44]
	0.42		[42]
Protective creams (sun, insects)	0.146	UV filters, glycol ethers, nanoparticles, insect repellents (e.g., N,N-diethyl-m-toluamide (DEET))	[45,46]
	0.090		[47]
Toothpaste	0.93	Antimicrobial agents, nanoparticles, microplastics	[44]
	1.00		[42]

Whereas personal-care products are linked primarily to bathing and hygiene, laundry-related products such as detergents and additives represent another major and chemically distinct emission pathway from everyday household activities. As the global population increases, demand for detergents and related additives is expected to rise accordingly [55]. Typical laundry detergents comprise complex formulations that may include surfactants, builders, alkalis, corrosion inhibitors, zeolites, processing aids, bleaching agents, colourants, fragrances, enzymes, and optical brighteners [56]. Depending on the product, additional auxiliaries such as defoamers or thickening agents may also be used. Detergents are marketed in multiple formats, including powders, liquids, and increasingly, single-dose capsules. Liquid detergents can contain up to ~50% surfactants by mass, whereas powder detergents generally contain substantially lower surfactant fractions.

Reported per-capita detergent consumption varies widely across regions and studies. Rotsidou et al. estimated a consumption of 1.99 kg person⁻¹ year⁻¹ in the UK [42], but substantially higher values have also been reported. For example, one Europe-wide estimate for 1998 reported annual consumption of 3.10 million tonnes of powder detergent, 0.56 million tonnes of liquid detergent, and 1.00 million tonnes of fabric softener. Using an estimated European population of 728 million, this corresponds to 4.25 kg of powder detergent, 0.77 kg of liquid detergent, and 1.37 kg of fabric softener per person per year (6.39 kg in total) [56]. Kruschwitz et al. reported even higher per-capita values, with 12.2 kg person⁻¹ year⁻¹ in 1984 and 7.15 kg person⁻¹ year⁻¹ in 2008 [57]. Such variation is consistent with differences in washing frequency, dose per wash, and geographically varying consumer behaviour. Minimum reported values include a washing frequency of 5.3 cycles per week and a dose of 64 g per wash. By contrast, US-EPA guidance indicates an average washing frequency of 1.32 cycles per day (9.24 per week) in the USA. Because the product dose can range from 20 to 290 g per wash (A.I.S.E.), Meesters et al. adopted 150 g per wash as a conservative representative value [58]. For Norway, a washing frequency of 8.3 cycles per week for a four-person household has been reported; assuming 150 g per wash, this yields an estimated detergent consumption of 16.1 kg person⁻¹ year⁻¹. Taken together, available evidence supports a global per-capita detergent consumption range of approximately 1.99–16.1 kg person⁻¹ year⁻¹, depending on usage patterns.

Stricter regulation of detergent formulations has reduced certain pollution pressures, notably through limitations on phosphorus-containing compounds [59]. Nevertheless, by mass, detergents remain a major household input to municipal wastewater systems. Recent studies further suggest that detergent degradation products may contribute to glyphosate burdens in receiving waters [10]. Thus, despite regulatory progress, the very large quantities used can influence wastewater treatment performance and contribute to the load of trace contaminants (e.g., phosphorus species, glyphosate, and surfactant transformation products).

In addition to laundry, routine household cleaning activities contribute further chemical loads that broaden both the composition and variability of domestic wastewater. For example, Meesters et al. reported typical use of an all-purpose cleaner at 65 g per application diluted into 5 L (13 g L⁻¹), applied 197 times per year, corresponding to 12.81 kg person⁻¹ year⁻¹ [59]. In practice, however, consumers use a range of products applied at different frequencies [60], including spray cleaners used in smaller volumes, such as kitchen and bathroom cleaners. These were reported to be applied at 22 g per use, with a frequency of 365 uses per year, yielding 8.03 kg person⁻¹ year⁻¹ [58]. By contrast, Rotsidou et al. reported substantially lower annual cleaning-agent consumption (1.57 kg person⁻¹ year⁻¹) [42]. Given the heterogeneity of products and application practices, and the limited availability of comparable datasets, it is reasonable to treat the reported range as an approximate estimate for per-capita annual use, while acknowledging substantial uncertainty. Car washing can occur at home or via commercial services. Reported washing frequencies of 11 events per year, with up to 100 g of product used per wash, correspond to an annual consumption of approximately 1.1 kg per person and year [61]. Overall, household and car-cleaning products can therefore be estimated at 2.67–12.91 kg person⁻¹ year⁻¹.

In addition to surfactants, builders, and fragrances, which have already been discussed as environmentally relevant, certain cleaning products (e.g., some toilet cleaners) have been reported to contain microplastics [62]. Garden maintenance introduces an additional class of potential pollutants through pesticide use. Misuse can cause documented adverse effects on consumer health. Pyrethroids such as tetramethrin, permethrin, and d-phenothrin are commonly used as domestic insecticides, while glyphosate-based formulations are widely used herbicides [63]. These substances can have negative environmental impacts; however, because reliable data on application quantities in the household context are scarce, no quantitative estimates are provided here.

A closely related, but sufficiently distinct, household activity is dishwashing, which introduces its own characteristic mixture of detergents, additives, and regeneration salts. Dishwashing may be performed manually using detergents or via dishwashers. In automated dishwashing, consumers typically use detergent tablets, rinse aids, and salt for regeneration of the machine’s built-in softener. The associated formulations contain surfactants, builders, alkalis, bleaching agents, and additives such as fragrances. Rotsidou et al. estimated 7.98 kg person⁻¹ year⁻¹ for manual dishwashing and 1.32 kg person⁻¹ year⁻¹ for automated dishwashing [42], giving a combined total of 9.30 kg person⁻¹ year⁻¹. Meesters et al. reported dishwashing frequencies of 252–365 events per year and product doses of 20–46 g per wash, corresponding to 16.79 kg year⁻¹ for the automated component under high-use assumptions [58]. They further reported manual dishwashing detergent use of 7 g per event at 426 events per year (2.98 kg year⁻¹). This yields a combined annual mass of 19.77 kg of dishwashing chemicals per household. Because household size was not specified, and given that these consumption patterns reflect predominantly Western settings, we normalised by an assumed mean household size of two [64], giving an estimated 9.89 kg person⁻¹ year⁻¹. This estimate closely matches the value reported by Rotsidou et al. While surfactant emissions are well recognised and regulated, emerging evidence attributes additional loads of phosphorus and microplastics to dishwashing-related products and practices [10,65].

Alongside product-derived chemical inputs, household wastewater also contains food-related residues that are highly relevant from an operational and environmental perspective. Fats, oils, and grease (FOG) do not necessarily constitute a chemical pollutant class in the conventional sense, but are included here because they can disrupt wastewater treatment processes and, critically, cause blockages and damage in sewer networks (including the formation of “fatbergs”) [66]. A key consequence is the increased likelihood of overflows and the discharge of untreated wastewater to the environment. Collin et al. discussed recovery of FOG from wastewater streams as a feedstock for biodiesel production, and reported domestic FOG inputs of 0.8 kg person⁻¹ year⁻¹ [67].

While FOG mainly affects conveyance and treatment performance, other household inputs are of concern because of their persistence in the environment, notably microplastics. Microplastic contamination has already been discussed for personal-care products and household cleaners. However, the dominant household source of microplastics entering domestic wastewater is laundry.

Table 3. Overview of microplastic emissions by private households into aquatic systems.

Entry Pathway	Total Emissions (Metric Tonnes)	Microplastic Emissions (Kg Person−1 Year−1)	Ref.
Personal care products (PCPs)	Toothpaste (India): 1400	Toothpaste (India): 0.00098	[68]
	Shower gel (China): 39	Shower gel (China): 0.000028	[68]
	Exfoliants (China): 307	Exfoliants (China): 0.00022	[68]
	Global PCP: 12,000	Global PCPs: 0.0016 (≈0.8% of total microplastics)	[69]
Dishwashing abrasion	Scouring pads (Denmark): 55	Scouring pads (Denmark): 0.0097	[70]
Laundry abrasion	1,500,000	0.201 (range: 0.0383–0.261)	[71]

summarises household-related microplastic emission pathways. Personal-care products are estimated to contribute 0.0012–0.0016 kg person⁻¹ year⁻¹ [68,69]. Abrasion during dishwashing, notably from scouring pads, has been estimated at 0.0097 kg person⁻¹ year⁻¹ [70]. By far the largest contribution is associated with laundering of textiles, estimated at 0.201 kg person⁻¹ year⁻¹ [71,72].

It should be noted that, in the context of laundry abrasion, “microplastics” are often quantified only as synthetic fibres. For natural fibres such as cotton (CO), complete biodegradability is frequently assumed [73]. Furthermore, while polyester (PES) shows poor biodegradability, a 50% CO/50% PES blend exhibits approximately 50% biodegradability (Figure 2). However, dyes and functional coatings can substantially affect degradability. For example, pink-dyed cotton shows reduced biodegradability and increased ecotoxicity relative to untreated cotton [73]. If coated or chemically modified natural fibres were included, the effective mass of persistent microfibre emissions would likely be higher. At present, however, the available evidence is insufficient to support robust quantitative estimates.

Beyond organic and polymer-based contaminants, domestic wastewater also receives inorganic pollutants that remain environmentally important despite their comparatively lower mass flows. An important member of this group are heavy metals, which, although present in trace amounts only in drinking water, have significant inputs from the disposal and use of household products such as cosmetics and cleaning agents [74], washing of workwear contaminated with metals [75], and leaching from damaged pipework and roof drainage systems [76]. Kalinowska et al. provide a useful overview of heavy-metal burdens in domestic wastewater, including a case study from Ostrava (Czech Republic) [77]. Summing reported metal concentrations yields a total of 3692.1 μg L⁻¹. Multiplying this concentration by the typical annual per-capita water use (43,800 L person⁻¹ year⁻¹) [34] gives an estimated annual heavy-metal load of domestic wastewater of 0.162 kg person⁻¹ year⁻¹.

A further low-mass but high-concern group comprises pharmaceuticals, whose environmental significance is driven less by quantity than by biological activity and persistence. These compounds are prescribed and consumed in large quantities, are excreted in human waste, and may also enter sewers through improper disposal via sinks or toilets, contributing to measurable concentrations in wastewater influent [76]. Estimates of the household contribution to overall antibiotic emissions vary by region and by compound. Dinh et al. reported that private households account for approximately 10% of total antibiotic emissions, with concentrations in household wastewater ranging from 0.03 to 1.75 μg L⁻¹ [77]. Using the annual per-capita water use (43,800 L person⁻¹ year⁻¹) [34], this corresponds to 1.3 × 10⁻⁶ to 7.7 × 10⁻⁵ kg person⁻¹ year⁻¹. Heberer et al. considered other pharmaceutical residues, including carbamazepine and diclofenac, and reported that 72.5% of emissions originated from private households. In that study, households emitted 1232.3 g week⁻¹ (64.23 kg year⁻¹) [78]. Normalising by the study population (96,000 inhabitants) yields 6.7 × 10⁻⁴ kg person⁻¹ year⁻¹. Summing the maxima across these studies gives an approximate upper-bound estimate of 7.5 × 10⁻⁴ kg person⁻¹ year⁻¹ for combined pharmaceutical-residue emissions. Although these masses are small, their environmental and public-health relevance is substantial because a fraction can pass through wastewater treatment and contribute to the spread of antimicrobial resistance. Illustratively, antibiotic-resistant Yersinia pestis strains have been isolated in Madagascar [79].

Taken together, these examples show that household wastewater contains both high-volume pollutant streams and low-concentration, high-impact contaminants, reinforcing the value of the comparative framework presented in Table 1. While reducing heavy-metal and pharmaceutical inputs is challenging and often requires upstream product redesign and improved medical and disposal practices, other high-mass inputs—such as chloride from softening salts or toilet paper consumption—may be more amenable to near-term reductions through targeted regulation and behavioural change.

2.2. Exposure Pathways in Wastewater Treatment Plants

Quantifying the mass of household-derived pollutants, and identifying where they enter the sewer network, is essential for designing effective regulation and for deploying targeted, innovative treatment technologies. However, the mass-based ranking in Table 1 does not capture how these pollutant groups behave within wastewater treatment plants (WWTPs). Some materials and components (e.g., cellulose fibres from toilet paper, and many readily biodegradable surfactants) are removed efficiently, whereas others—most notably PFAS and many pharmaceutical residues—are persistent and therefore far more challenging. In principle, WWTPs can be retrofitted with advanced (often costly) polishing steps, such as activated carbon adsorption, which can substantially reduce emissions of specific compounds (including PFAS) to the environment [80]. In practice, these upgrades are not implemented at scale worldwide because of the associated costs. Similarly, integrating additional treatment modules into septic-tank systems remains an active area of research. For this reason, current and emerging technology options for mitigating pollutant emissions from domestic wastewater are discussed separately in Section 3. Here, we focus on the present global situation and the performance of conventional treatment trains.

Globally, only 56% of wastewater is treated safely [81], typically interpreted as receiving at least secondary treatment (i.e., treatment beyond preliminary and primary steps). The extent of tertiary treatment varies widely between regions and even within the USA and the EU. Data from the European Environment Agency indicate that in many EU countries, tertiary treatment is still only partially implemented [82]. By contrast, “quaternary” or advanced treatment (e.g., adsorption, ozonation, and other specialised processes) is applied only in a limited number of cases.

A substantial fraction of wastewater worldwide is treated through decentralised, on-site systems, most commonly septic tanks. Septic tanks are not optimised for highly loaded household wastewater, and their performance depends strongly on design. Nasr et al. compared conventional septic tanks, single- and two-chamber systems, and fixed-bed (attached-growth) configurations, and reported that two-chamber systems and fixed-bed septic tanks provide the most robust treatment for high-strength domestic wastewater [83]. Even then, additional post-treatment is generally required to address specific pollutants. Septic systems are also sensitive to influent characteristics. For example, fats, oils, and grease (FOG) and elevated salinity can impair system performance. In cold seasons in particular, FOG can solidify and cause clogging and operational failures; grease interceptors are therefore needed, but are often not designed or operated to handle variable household FOG loads reliably, which can lead to process disturbances [84]. High chloride concentrations, for example due to the use of water softeners, can further reduce septic-tank performance [31]. Pawlak et al. reported that operation of water softeners can increase chloride concentrations in septic tanks by up to 4.5 g L⁻¹ [31], levels that may disrupt downstream biological degradation processes. Given the global relevance of decentralised treatment—approximately 23% of the 115 million houses in the USA are connected to on-site systems [84]—a sound understanding of domestic wastewater composition is required both to avoid process disruptions and to design fit-for-purpose treatment solutions. For many trace contaminants, including PFAS, polycyclic aromatic hydrocarbons (PAHs), and antibiotics, more advanced treatment is typically required.

We now turn to conventional municipal WWTPs and the behaviour of major household-related pollutant groups. Salinity inputs, primarily as sodium chloride, are particularly relevant because NaCl is not retained by standard treatment processes. Removing salt generally requires energy-intensive membrane processes (e.g., reverse osmosis) together with effective pretreatment to manage fouling and concentrate handling. As in decentralised systems, elevated salinity can impair biological performance once certain thresholds are exceeded. Jovanović et al. reported inhibition of nitrification at chloride concentrations above 1 g L⁻¹, whereas below this threshold, moderate salinity may increase nitrification rates [85]. Lu et al. systematically investigated the effects of increasing chloride on organic-matter removal and nitrogen and phosphorus removal in sequential batch reactors (SBRs) and the A²/O (anaerobic–anoxic–oxic) process. When chloride was increased from 150 to 5000 mg L⁻¹, COD removal decreased by 48% and 31.8% for the two reactor configurations, and nitrification and phosphorus removal were also inhibited [86]. Together, these studies illustrate how high salinity can compromise key biological treatment functions.

Toilet paper primarily enters WWTPs as a particulate load that breaks down into cellulose fibres. These fibres are largely removed during mechanical pretreatment (screens and sieves). Nevertheless, life-cycle assessment (LCA) studies indicate that removal of toilet-paper residues can be energy-intensive. Even so, installing screening units was associated with a 9.46% reduction in overall environmental impact in the assessed system [87]. These findings suggest that further improving mechanical pretreatment and recovering cellulose-rich residues could benefit both WWTP performance and the wider environmental footprint [87]. Importantly, while the bulk cellulose fraction is removed, toilet paper can also act as a carrier of chemical contaminants, including PFAS, which are poorly removed by conventional treatment. A detailed Dutch study measured PFAS in influent, effluent, and sludge at eight conventional WWTPs [88]. Total PFAS concentrations in influent and effluent ranged from 10 to 1000 ng L⁻¹, with limited removal and, in some cases, higher concentrations detected in the effluent than in the influent [88]. Most PFAS left the plants via the effluent, predominantly as short-chain PFAS, whereas long-chain PFAS and precursor compounds were more strongly associated with sewage sludge, with concentrations of 10–100 μg kg⁻¹ dry mass [88]. The persistence and partitioning behaviour of PFAS therefore represent a substantial environmental concern.

Laundry detergents, dishwashing products, household cleaners, and associated additives (e.g., fabric softeners and builder components) are ubiquitous in domestic wastewater and can, under certain conditions, challenge WWTP operation. Wastewater dominated by food- and cleaning-related inputs is commonly classified as greywater and shows characteristic physicochemical properties [89]. Liquid detergents and cleaners typically contribute more strongly to COD and total organic carbon (TOC), but can contain lower fractions of suspended solids and linear alkylbenzene sulfonates (LAS) than powder-based formulations. Interestingly, coagulation–flocculation, which is widely used for solids separation, has been reported to perform better for wastewater dominated by powder detergents than for liquid-detergent wastewater [90]. Such differences highlight the importance of understanding source-specific wastewater streams when selecting or optimising treatment processes [90].

Despite regulatory progress, detergents and cleaners can still contribute materially to phosphorus loads. Richards et al. compared phosphorus contents in conventional versus “eco” products. Conventional dishwasher detergents showed the highest phosphorus contents (43–131 mg P g⁻¹ detergent), whereas eco-labelled dishwasher products had substantially lower values (0.7–9.1 mg P g⁻¹ detergent) [91]. Conventional dishwasher detergents accounted for approximately 95% of the phosphorus input from all washing and cleaning products [91]. High phosphorus loads complicate wastewater treatment and can increase costs, for example through higher sludge production and interference with downstream biological processes, motivating ongoing research into cost-effective mitigation strategies [91]. Regulatory discharge limits may be exceeded in older or poorly controlled WWTPs, and in septic systems. Particular attention has recently focused on aminopolyphosphonates used as builder components, which can transform to glyphosate in sludge under specific conditions, including in the presence of manganese compounds [10,92]. These findings underscore that part of the glyphosate burden in receiving waters may originate from detergent additives [10,92].

Pesticides are also regularly detected in domestic wastewater and urban drainage, including inputs associated with green-space maintenance in residential areas [93]. Influent concentrations for most pesticides are typically below 1 μg L⁻¹, yet removal in conventional WWTPs is often poor [94]. Process optimisation, for example by enhancing sorption, has been proposed to improve removal performance [95]. In some cases, higher concentrations have been reported in effluent than in influent, which may reflect transformation processes (e.g., deconjugation or release from particulate matter) during treatment rather than true “generation” in the plant [94].

Pharmaceuticals and personal care products (PPCPs) are often considered together because many are persistent organic micropollutants that are incompletely removed by conventional treatment. Primary and secondary treatment can remove up to ~90% for certain compounds, but achieving very low effluent concentrations commonly requires additional processes such as ozonation or membrane filtration [96]. Focusing on personal-care products, synthetic fragrances and UV filters (particularly during warmer seasons) can account for a substantial fraction of influent loads. Biel-Maeso et al. reported comparatively low removal for several personal-care compounds, with removal efficiencies below 70% [97]. Pharmaceuticals, including antibiotics, are likewise challenging, and their presence raises additional concerns regarding antimicrobial resistance. A study of three WWTPs in Romania reported substantial antibiotic concentrations in effluent, with seasonal variability and higher concentrations in warmer months. Resistance to antibiotics—particularly amoxicillin and clarithromycin—also increased during warmer periods [98]. Overall, available evidence indicates that PPCPs remain a persistent challenge and that continuous emission to the environment occurs under widely deployed treatment configurations.

Microplastics can also disrupt treatment performance and are incompletely removed in many systems. Summarising microplastic detection across different treatment stages, Sadia et al. concluded that plants limited to primary and secondary treatment often show comparatively low microplastic removal efficiencies. WWTPs are therefore considered a major pathway for microplastic release to aquatic environments. High removal efficiencies (up to 99.4%) have been reported for membrane bioreactors (MBRs) [99]. However, membrane fouling can compromise filtration performance, indicating that further development and optimisation of membrane technologies remain important [100].

Many WWTPs are also not designed for substantial heavy-metal removal. Metals are commonly detected both in effluent and in sewage sludge. Domestic wastewater can contribute strongly to WWTP metal loads, particularly during dry periods when household wastewater constitutes a larger fraction of total inflow. For the Ostrava WWTP (Czech Republic), influent source apportionment suggested that domestic wastewater accounted for ~60% of Cd, 21–35% of Zn, Cu, and As, and <15% of Pb, Cr, Fe, Ni, and Mn [101]. Managing these metals remains challenging and is an active research area. Cost-effective strategies are required, not least to enable environmentally sound reuse of sewage sludge in agriculture. Coagulation–flocculation and sorption processes in early treatment stages have been identified as promising approaches for reducing metal concentrations [102].

Overall, the evidence reviewed above shows that pollutant mass alone does not determine treatment challenge; rather, persistence, partitioning behaviour, and process sensitivity are equally important in shaping environmental release. Section 2.2 highlighted that the pollutant groups ranked in Table 1 pose distinct and, in many cases, significant challenges for wastewater treatment. Given that only 56% of wastewater is treated safely worldwide [81], emissions of household-derived pollutants to the environment remain widespread.

2.3. Environmental and Societal Challenges Associated with Emissions of Domestic Wastewater Pollutants

As noted above, only a fraction of global wastewater is treated safely. Moreover, even where “safe treatment” is reported, this often refers to conventional WWTP configurations that are not designed or optimised to remove the broad range of household-derived pollutants now recognised as environmentally relevant. This gap has direct societal implications, including rising costs for monitoring, treatment upgrades, and remediation. Table 4 summarises key environmental and societal impacts associated with major household-related pollutant groups and provides illustrative examples of countermeasures. Some of these countermeasures focus on changes in consumer behaviour (e.g., awareness campaigns), while the technological options listed are restricted to measures that are likely to be cost-effective at scale. Retrofitting WWTPs or decentralised on-site systems with advanced processes (e.g., adsorption, ozonation) is technically feasible but remains complex and is an active research area (cf., Section 3).

Table 4 provides an overview of the impacts of household-derived pollutants. It is important to recognise that identifying sources, understanding environmental partitioning and accumulation, and evaluating toxicity and risk are areas of ongoing research and key uncertainties persist; hence Table 4 should not be considered a definitive or exhaustive listing. For example, salinisation of rivers and lakes is an increasing global concern [103,104]. Because chloride from wastewater inputs (e.g., road salt, water softeners) is problematic for WWTPs and septic systems [31,85] and is not removed by conventional treatment, it ultimately reaches surface waters [105], with documented negative effects on aquatic organisms and ecosystem functioning [106,107]. Reducing or avoiding salt-based water softening is therefore seen as a possible practical option to lower chloride loads [24].

This broader challenge is not limited to salinisation, but also extends to persistent organic pollutants and particulate contaminants for which sources, pathways, and risks remain only partly resolved. For instance, PFAS concentrations in precipitation have increased substantially over the past ~25 years, by factors of approximately 4–6, while only about 60–70% of sources can be attributed to known pathways [108]. Improving the identification and quantification of PFAS entry pathways is therefore critical, yet comprehensive risk assessment remains incomplete. For microplastics, more integrated assessment concepts have been developed in recent years. Figure 3 illustrates an ecotoxicological screening approach aligned with DIN SPEC 4872. While synthetic fibres represent the dominant concern in many contexts, coated or dyed cotton fibres can also contribute to microfibre emissions with reduced biodegradability and measurable ecotoxicological effects [73]. Together, these examples underline that substantial open questions remain regarding the sources, environmental distribution, and impacts of domestic wastewater pollutants. Addressing these questions through robust monitoring and targeted research is essential for enabling effective, evidence-based regulation at scale.

Table 4. Overview of environmental and societal impacts of domestic wastewater pollutant groups and illustrative countermeasures.

Environmental and Societal Impacts	Illustrative Countermeasures	Ref.
Sodium chloride: Freshwater salinisation; harm to aquatic organisms and biodiversity; impaired ecosystem functioning; reduced drinking-water quality; infrastructure corrosion; potential health risks from elevated salt intake.	Consider centralised softening where appropriate; promote salt-free softening technologies; restrict excessive road-salt use, including private use; develop less harmful de-icing alternatives.	[25,31,36,106,107]
PFAS: Persistent and bioaccumulative; transport in water and sediments; toxicity to aquatic organisms; potential human-health risks.	Improve monitoring and risk assessment; reduce household PFAS sources; consider multifunctional toilets where LCA supports benefits; promote PFAS alternatives in consumer products and textiles; strengthen regulation, including drinking-water limits.	[109,110,111]
Surfactants: Environmental occurrence despite biodegradability; potentially harmful transformation products; endocrine-disrupting activity for some compounds; toxicity to aquatic and terrestrial organisms.	Promote eco-labelled products; strengthen monitoring and regulation of surfactants and their transformation products; where no WWTP connection exists, ensure effective septic treatment.	[112,113,114]
Phosphates/Phosphonates: Eutrophication; algal and cyanobacterial blooms; oxygen depletion; ecological degradation of receiving waters; possible transformation of phosphonates to glyphosate; operational problems in WWTPs.	Improve phosphorous removal and recovery; encourage lower-phosphorus products; tighten regulation of phosphonates, particularly in dishwashing detergents; improve septic-system management where sewer connection is absent.	[10,58,114,115]
Endocrine disruptors in PCPs: Persistent, bioactive, and in some cases bioaccumulative; compounds such as triclosan, parabens, preservatives, and UV filters can disrupt aquatic systems; additional microplastic emissions.	Strengthen monitoring and regulation of PCPs; improve consumer awareness and product labelling; enhance trace-level detection and risk assessment	[116,117,118]
Microplastics: Adverse effects on flora and fauna; accumulation in sediments; transport vector for co-contaminants (e.g., PFAS); cellular toxicity after uptake.	Enforce restrictions on microplastics in personal-care products; encourage lower-impact consumer choices; improve textile abrasion resistance and washing-machine design to reduce fibre release.	[119,120,121,122]
Heavy metals: Persistent, bioaccumulative, and often toxic or carcinogenic; some metals pose risks at trace levels; ecotoxic stress in aquatic systems.	Improve household source monitoring; support lower-impact product choices through consumer information; develop effective removal methods for WWTPs and septic systems, including biological options where appropriate.	[74,123,124,125]
Pesticides: Low target efficiency with substantial environmental release; household wastewater can contribute to pesticide loads; some compounds show endocrine-disrupting potential; chronic effects remain uncertain.	Tighten regulation of households and residential pesticide use; strengthen monitoring and risk assessment; re-evaluate household wastewater as a potential glyphosate source; promote biopesticides where effective and environmentally preferable.	[10,93,126]
Pharmaceuticals: Persistent or transformed into products with uncertain effects; risks to aquatic and terrestrial organisms; chronic human exposure via water and food; endocrine effects and antimicrobial resistance with major societal costs.	Promote green and sustainable pharmacy; strengthen source-specific monitoring and risk assessment; improve public education on appropriate medicine use and disposal; integrate environmental criteria into healthcare guidance; develop cost-effective WWTP upgrades; tighten regulations and limit values.	[127,128]

Abbreviations: LCA, Life-cycle assessment; PCPs, personal care products; PFAS, perfluoroalkyl and polyfluoroalkyl substances; WWTP, wastewater treatment plants.

Table 4. Overview of environmental and societal impacts of domestic wastewater pollutant groups and illustrative countermeasures.

Environmental and Societal Impacts	Illustrative Countermeasures	Ref.
Sodium chloride: Freshwater salinisation; harm to aquatic organisms and biodiversity; impaired ecosystem functioning; reduced drinking-water quality; infrastructure corrosion; potential health risks from elevated salt intake.	Consider centralised softening where appropriate; promote salt-free softening technologies; restrict excessive road-salt use, including private use; develop less harmful de-icing alternatives.	[25,31,36,106,107]
PFAS: Persistent and bioaccumulative; transport in water and sediments; toxicity to aquatic organisms; potential human-health risks.	Improve monitoring and risk assessment; reduce household PFAS sources; consider multifunctional toilets where LCA supports benefits; promote PFAS alternatives in consumer products and textiles; strengthen regulation, including drinking-water limits.	[109,110,111]
Surfactants: Environmental occurrence despite biodegradability; potentially harmful transformation products; endocrine-disrupting activity for some compounds; toxicity to aquatic and terrestrial organisms.	Promote eco-labelled products; strengthen monitoring and regulation of surfactants and their transformation products; where no WWTP connection exists, ensure effective septic treatment.	[112,113,114]
Phosphates/Phosphonates: Eutrophication; algal and cyanobacterial blooms; oxygen depletion; ecological degradation of receiving waters; possible transformation of phosphonates to glyphosate; operational problems in WWTPs.	Improve phosphorous removal and recovery; encourage lower-phosphorus products; tighten regulation of phosphonates, particularly in dishwashing detergents; improve septic-system management where sewer connection is absent.	[10,58,114,115]
Endocrine disruptors in PCPs: Persistent, bioactive, and in some cases bioaccumulative; compounds such as triclosan, parabens, preservatives, and UV filters can disrupt aquatic systems; additional microplastic emissions.	Strengthen monitoring and regulation of PCPs; improve consumer awareness and product labelling; enhance trace-level detection and risk assessment	[116,117,118]
Microplastics: Adverse effects on flora and fauna; accumulation in sediments; transport vector for co-contaminants (e.g., PFAS); cellular toxicity after uptake.	Enforce restrictions on microplastics in personal-care products; encourage lower-impact consumer choices; improve textile abrasion resistance and washing-machine design to reduce fibre release.	[119,120,121,122]
Heavy metals: Persistent, bioaccumulative, and often toxic or carcinogenic; some metals pose risks at trace levels; ecotoxic stress in aquatic systems.	Improve household source monitoring; support lower-impact product choices through consumer information; develop effective removal methods for WWTPs and septic systems, including biological options where appropriate.	[74,123,124,125]
Pesticides: Low target efficiency with substantial environmental release; household wastewater can contribute to pesticide loads; some compounds show endocrine-disrupting potential; chronic effects remain uncertain.	Tighten regulation of households and residential pesticide use; strengthen monitoring and risk assessment; re-evaluate household wastewater as a potential glyphosate source; promote biopesticides where effective and environmentally preferable.	[10,93,126]
Pharmaceuticals: Persistent or transformed into products with uncertain effects; risks to aquatic and terrestrial organisms; chronic human exposure via water and food; endocrine effects and antimicrobial resistance with major societal costs.	Promote green and sustainable pharmacy; strengthen source-specific monitoring and risk assessment; improve public education on appropriate medicine use and disposal; integrate environmental criteria into healthcare guidance; develop cost-effective WWTP upgrades; tighten regulations and limit values.	[127,128]

Abbreviations: LCA, Life-cycle assessment; PCPs, personal care products; PFAS, perfluoroalkyl and polyfluoroalkyl substances; WWTP, wastewater treatment plants.

3. Treatment Technologies for Domestic Wastewater Pollutants

Conventional approaches for the removal of emerging organic contaminants from domestic wastewater—including cleaning and detergent agents, antibiotics, microplastics, and other micropollutants (e.g., parabens, preservatives, hormones, ultraviolet (UV) filters, and antioxidants)—primarily comprise classical filtration (e.g., granular media filtration and cartridge filters) and low-pressure membrane processes, notably microfiltration (MF) and ultrafiltration (UF). However, these methods are generally ineffective for dissolved organic contaminants, and, in practice, provide efficient removal mainly for the microplastic fraction with particle sizes above approximately 1 μm. Among low-pressure membranes, UF is the most capable of retaining submicrometre particles and can achieve high removal efficiencies for microplastic particles smaller than 1 µm. Nevertheless, because these processes are operationally simple and comparatively inexpensive, they are widely implemented as preliminary or prefiltration steps in water and wastewater treatment.

A second group of technologies comprises high-pressure membrane processes, which can achieve high removal efficiencies for both dissolved organic contaminants and microplastics. Their principal drawbacks are high energy demand, alongside substantial capital and operating costs. Despite these limitations, the high removal performance of these processes warrants their inclusion in this review. In addition, membrane separation is frequently integrated with biological treatment in membrane bioreactor (MBR) configurations. MBRs combine high biological activity with a physical membrane barrier, enabling the retention and biodegradation of a broad spectrum of organic contaminants, alongside efficient removal of microplastic particles.

Despite progress in conventional treatment and membrane-based processes, the reliable removal of emerging organic contaminants remains a major challenge. Many such compounds exhibit high chemical stability, aromatic or heterocyclic structures, and multiple functional groups, which often translate into low biodegradability and poor removal by standard biological and physicochemical processes. Importantly, many of these contaminants share structural features—including aromatic rings, conjugated π-electron systems, and polar functional groups—that govern both their fate and their amenability to treatment. These moieties can promote adsorption via π–π interactions, hydrogen bonding, and electrostatic forces, facilitating preconcentration of pollutants on functional surfaces in adsorption-based processes. At the same time, the same structural elements are often susceptible to oxidative attack by reactive oxygen species generated in photocatalytic processes, enabling efficient degradation and, in some cases, extensive mineralization. In this context, advanced oxidation processes (AOPs), particularly heterogeneous photocatalysis, have attracted increasing attention as promising approaches for degrading persistent organic pollutants.

Among material classes with high potential for both adsorption and photocatalysis are metal–organic frameworks (MOFs), covalent organic frameworks (COFs), and hybrid MOF–COF structures. As discussed in subsequent sections, these materials can function as highly efficient adsorbents, as porous photocatalysts, or as functional additives for modifying filtration and membrane materials, thereby enhancing both separation performance and contaminant degradation.

3.1. Filtration Technologies

This subsection focuses on granular media filtration and low-pressure membrane processes (MF and UF), which are widely used as prefiltration steps in water and wastewater treatment. Owing to their pore size characteristics, these processes are generally ineffective at removing dissolved emerging organic contaminants—such as cleaning and detergent agents, antibiotics, and other micropollutants—by size exclusion, and any removal is typically limited to non-specific adsorption with low practical capacity [129,130,131].

By contrast, these filtration technologies can be effective for removing microplastic particles. Literature reports indicate that granular filtration using different sand-filter configurations can remove microplastics >1 μm with efficiencies exceeding 95% [132]. Carbon-based filtration has also been reported to achieve relatively high removal efficiencies (73.7–98.5%) for small microplastic particles in the 1–5 μm range [133]. In addition, polypropylene filters have been reported to remove >95% of microplastic particles with sizes between 1 and 2 μm [134]. These findings indicate that while conventional filtration is inadequate for most dissolved emerging organic contaminants, it remains a useful tool for the physical removal of microplastics within specific size ranges. MF [132,135,136] and UF [132,137,138,139] are likewise characterised by high microplastic removal efficiencies. For UF, near-complete removal of very small particles has been reported under certain operating conditions, including for the nanoplastic size range [140,141,142].

3.2. Nanofiltration and Reverse Osmosis

Nanofiltration (NF) and reverse osmosis (RO) are among the few treatment technologies capable of substantially reducing emissions of key contaminant classes relevant to domestic wastewater, including per- and polyfluoroalkyl substances (PFAS), antibiotics, microplastics, and inorganic salts. Their application addresses the limited performance of conventional biological and mechanical treatment with respect to persistent, mobile, and biologically active pollutants.

PFAS are among the most challenging contaminant groups in domestic wastewater because of their exceptional chemical stability, resistance to biodegradation, and high environmental mobility. In conventional WWTPs, PFAS largely pass through treatment, while long-chain PFAS may partially accumulate in sewage sludge. NF can retain long-chain PFAS, primarily via a combination of size exclusion and electrostatic interactions [143,144], but removal of short-chain PFAS is often lower. NF is therefore frequently considered an intermediate, load-reduction step prior to further advanced treatment [145]. RO is currently regarded as one of the most effective barrier technologies for PFAS removal from domestic wastewater [146]. RO can provide very high retention of both long- and short-chain PFAS, supporting applications aimed at minimising PFAS discharges to surface waters or producing high-quality reclaimed water. However, RO does not destroy PFAS; instead, it concentrates them in the retentate, which requires appropriate downstream management and disposal or destruction technologies [147].

Antibiotics in domestic wastewater, even at trace concentrations, pose environmental risks because of their role in selecting for antibiotic-resistant bacteria and promoting the dissemination of resistance genes. NF can remove many antibiotics with medium to high molecular weight, with separation governed by size exclusion and electrostatic interactions between ionisable pharmaceutical compounds and the charged membrane surface [148]. RO generally provides the highest removal performance for antibiotics and other pharmaceutical residues and is therefore commonly used as a polishing step in treatment trains designed to minimise releases of bioactive compounds. As with PFAS, accumulation of pharmaceuticals in the concentrate stream remains a critical practical issue.

Microplastics in domestic wastewater occur across a wide range of sizes and morphologies, including textile fibres and fragmented particles. While a substantial fraction is removed during mechanical and biological treatment, fine particles—particularly those <1 μm—can pass through conventional stages. Both NF and RO act as highly effective physical barriers for microplastics across particle sizes, shapes, and polymer types [149,150]. In practice, NF and RO can therefore function as safeguard technologies to limit microplastic release in treated effluent. This role is especially relevant in multi-barrier systems, including MBR-based configurations, where MF/UF stabilise feed-water quality for downstream high-pressure membranes and reduce fouling risks while supporting high overall particle retention [150].

Salinisation of domestic wastewater, driven by salt-based water softening and the seasonal application of road de-icing salts, is a growing challenge for WWTP operation and receiving waters. NF has limited capacity to remove monovalent ions such as Na⁺ and Cl⁻, and is therefore not well suited to direct salinity reduction. NF may, however, influence ionic composition indirectly by preferentially removing multivalent ions and some complexing species. By contrast, RO remains the membrane technology most capable of removing monovalent salts effectively from wastewater. RO is therefore increasingly considered in regions affected by salinisation of surface waters, or where effluent-quality requirements impose strict limits on conductivity and chloride concentrations. In this context, RO can contribute both to water-quality improvement and to mitigation of anthropogenic pressures that may be compounded by climate-related changes in hydrology and dilution capacity [140].

Overall, NF and RO play distinct yet complementary roles in advanced domestic wastewater treatment. NF is typically applied as a load-reducing step for organic micropollutants and selected PFAS, whereas RO functions as a final, high-performance barrier for PFAS, antibiotics, microplastics, and inorganic salts. In most practical applications, integration with biological treatment and/or additional advanced processes is essential to achieve sustained reductions in environmental impacts associated with domestic wastewater discharges.

3.3. Membrane Bioreactor

Membrane bioreactors (MBRs) can achieve high removal efficiencies for microplastics and many organic contaminants by combining biological degradation with physical separation in a single, integrated treatment train [151]. MBR configurations are also technologically flexible because they can be implemented with different membrane processes and operational modes. Performance can be further improved by coupling MBRs with advanced oxidation processes (AOPs), such as ozonation, UV/H₂O₂, or photocatalysis, and by hybridising MBRs with electrochemical processes, including electrocoagulation and electro-oxidation [152,153,154]. Reported microplastic removal efficiencies in MBR systems commonly exceed 99% [155,156,157]. Removal rates above 99% have been reported for microplastic particles of approximately 0.25 µm [158], and high retention has also been reported for particles in the nanoplastic size range [159].

A major limitation of MBR operation is membrane fouling and biofouling, which can reduce flux, increase energy demand, and ultimately lower overall process performance. In addition, microplastics may influence the removal of other contaminants by affecting microbial activity and enzyme-mediated processes within the bioreactor [160,161,162]. For organic contaminants, removal efficiencies depend strongly on compound-specific physicochemical properties (e.g., polarity, charge state, and sorption behaviour). Accordingly, within a single MBR system, some substances may be removed at >90%, whereas others exhibit substantially lower removal [163,164,165]. Nevertheless, for a broad range of compounds, MBR systems have been reported to achieve removal efficiencies above 90% [166,167,168].

3.4. Adsorption Technologies

Although full-scale data on adsorption for the removal of microplastics, antibiotics, pharmaceuticals and personal care products (PPCPs), and softening and cleaning agents remain limited, extensive laboratory studies are evaluating novel adsorbents. These studies aim to elucidate adsorption mechanisms and to develop scalable synthesis routes for high-performance materials. Continued progress in this area may improve the efficiency and competitiveness of adsorption relative to alternative treatment technologies. The development of new sorbents has focused not only on increasing accessible surface area, but also on surface functionalisation to enhance selectivity and overall adsorption performance.

Metal–organic frameworks (MOFs), covalent organic frameworks (COFs), and hybrid MOF–COF structures are particularly promising for adsorption of organic contaminants because they can combine very high specific surface area, ordered porosity, and high thermal and chemical stability [169,170,171]. Representative applications of these structures are summarised in Table 5. In addition, these materials offer chemically tunable pore environments and surface properties, enabled by the incorporation of different functional groups. Functionalisation with carboxyl, aldehyde, hydroxyl, or amino groups, and/or aromatic moieties can enhance adsorption selectivity towards specific organic compounds while maintaining high accessible surface area [172,173]. MOF- and COF-based materials have also shown high affinity towards PFAS in adsorption studies [174,175].

Compared with conventional adsorbents, MOFs and COFs provide a distinctive combination of ordered porosity and programmable surface chemistry, allowing adsorption capacity and selectivity to be optimised in parallel. Possible interaction mechanisms with target contaminants include π–π stacking, electrostatic interactions, acid–base interactions, hydrophobic interactions, hydrogen bonding, and, in some cases, covalent bonding [176,177,178,179,180].

Table 5. Adsorption data for microplastics (MP) and selected organic contaminants (pharmaceuticals, hormones, pesticides, and dyes) reported for MOF, COF, and hybrid MOF–COF materials.

Material	Contaminant(s)	Removal Efficiency	Proposed Interaction(s)	Ref
TAPB-TPA (COF)	Polystyrene (MP)	84–95%	Electrostatic interactions	[181]
ZIF-67 (MOF)	Polystyrene (MP)	92.1%	Hydrogen bond interactions, π–π stacking, electrostatic interactions	[182]
COF(MATPA)-MOF(Zr)	Difenoconazole; tetrafluranazole	89.6%; 85.1%	-	[183]
CCF@UiO-66-NH2@TpBD	Bisphenol	~90%	π–π stacking, hydrogen bonding, hydrophobic interactions	[184]
MIL-53-C (MOF)	Methyltestosterone; testosterone propionate; nandrolone phenylpropionate	~80%; ~70%; ~60%	-	[185]
COF-SO3H	Indomethacin; diclofenac; ketoprofen	95%; 94%; 57%	hydrogen bonding, π–π stacking	[186]
MOF-5/COF (M5C)	Auramine O, rhodamine B (dyes)	~98%	π–π stacking, electrostatic interactions	[187]
TpStb-SO3	Methylene blue; crystal violet; malachite green; Janus green	>99	π–π stacking	[188]
ZIF-8 + TpPa (MOF–COF)	Tetracycline	62%	hydrogen bonding, π–π stacking	[189]

Table 5. Adsorption data for microplastics (MP) and selected organic contaminants (pharmaceuticals, hormones, pesticides, and dyes) reported for MOF, COF, and hybrid MOF–COF materials.

Material	Contaminant(s)	Removal Efficiency	Proposed Interaction(s)	Ref
TAPB-TPA (COF)	Polystyrene (MP)	84–95%	Electrostatic interactions	[181]
ZIF-67 (MOF)	Polystyrene (MP)	92.1%	Hydrogen bond interactions, π–π stacking, electrostatic interactions	[182]
COF(MATPA)-MOF(Zr)	Difenoconazole; tetrafluranazole	89.6%; 85.1%	-	[183]
CCF@UiO-66-NH2@TpBD	Bisphenol	~90%	π–π stacking, hydrogen bonding, hydrophobic interactions	[184]
MIL-53-C (MOF)	Methyltestosterone; testosterone propionate; nandrolone phenylpropionate	~80%; ~70%; ~60%	-	[185]
COF-SO3H	Indomethacin; diclofenac; ketoprofen	95%; 94%; 57%	hydrogen bonding, π–π stacking	[186]
MOF-5/COF (M5C)	Auramine O, rhodamine B (dyes)	~98%	π–π stacking, electrostatic interactions	[187]
TpStb-SO3	Methylene blue; crystal violet; malachite green; Janus green	>99	π–π stacking	[188]
ZIF-8 + TpPa (MOF–COF)	Tetracycline	62%	hydrogen bonding, π–π stacking	[189]

3.5. Advanced Oxidation Processes: MOF/COF Heterojunction Photocatalysts

Advanced oxidation processes (AOPs) are among the most promising approaches for removing persistent organic pollutants and, in some cases, contributing to the degradation of microplastic-associated contaminants [190,191,192]. Core AOP techniques include UV–hydrogen peroxide (UV/H₂O₂), Fenton and photo-Fenton reactions, ozone-based processes, and photocatalysis [193,194]. Photocatalysis is particularly relevant for persistent organic pollutants because reactive oxygen species can be generated in situ under mild conditions, with comparatively low chemical input. The process is often described as environmentally favourable because it uses light energy to drive oxidative transformations [195,196].

A key research focus is the design and evaluation of photocatalysts, including metal oxides (e.g., TiO₂ and ZnO), modified and doped semiconductors, sulphide- and nitride-based photocatalysts (e.g., CdS and g-C₃N₄), and composite or heterojunction systems intended to enhance charge separation and visible-light activity [197,198,199]. In this context, hybrid heterojunction architectures based on MOF–COF systems have attracted increasing attention because they can integrate visible-light-active COFs with the high porosity, large specific surface area, adsorption capacity, and catalytically active metal centres of MOFs. Formation of an intimate MOF–COF heterojunction—often described in terms of a Z-scheme charge-transfer pathway—can promote charge separation, suppress electron–hole recombination, and enhance reactive oxygen species generation [200,201,202]. Representative MOF–COF photocatalysts and their reported performance are summarised in

Table 6. Photocatalytic performance of reported MOF–COF materials towards selected organic contaminants (pharmaceuticals, parabens, and dyes).

Material	Contaminant(s)	Removal Efficiency	Ref
TA-BPDA-COF@ZIF-L-Co	Rhodamine B;	>98%;	[202]
	methyl orange;	>98%;
	methyl blue;	>98%;
	bisphenol A	>90%
NH2-MIL-125@TpMA	Methylparaben; ethylparaben;	86.21%;	[203]
	propylparaben	92.44%;
	(Parabens)	98.84%
NH2-MIL-125(Ti)@SNW-1	Tetracycline	61.37%	[204]
NH2-MOF-5/MCOF	Methyl blue	~90%	[205]
H2-MIL-125(Ti)@TAPB-TBAB	Tetracycline;	83%;	[206]
	carbamazepine;	65%;
	methyl orange;	96%;
	levofloxacin	58%
MIL-68@COF-V	Tetracycline; rhodamine; phenol	~96.5%; ~97.6%; ~95.3%	[207]
COF/MIL100/CuFe2O4	Malachite green;	98%;	[208]
	tetracycline	91%

.

3.6. Electrochemical Treatment Methods

Electrochemical treatment encompasses a range of technologies used to remove organic matter (including trace organic contaminants) and, indirectly, to facilitate the removal of (heavy) metals from wastewater. Electro-oxidation (EO), which can be classified as an AOP (see Section 3.5), generates hydroxyl radicals at the electrode surface (e.g., boron-doped diamond, mixed-metal oxides, or noble-metal electrodes) and typically requires only electrical energy rather than added oxidants. By degrading complexing agents (e.g., EDTA), EO may also promote precipitation of metals, thereby facilitating subsequent separation. Because electrode performance is central to EO, substantial development is ongoing, including the commercialisation of structured boron-doped diamond (BDD) electrodes supported on electrically insulating ceramics [209]. Designs with very small interelectrode gaps (<50 µm) can enable EO operation in low-conductivity electrolytes, potentially reducing or avoiding the need to add supporting salts. To the best of our knowledge, however, EO (with or without BDD) is not yet widely implemented in large-scale municipal WWTPs.

Other electrochemical methods include electrocoagulation (EC) and electroflotation (EF). These approaches are commonly used in small-scale or specialised facilities, particularly for industrial wastewaters [210]. In contrast, municipal WWTPs more commonly apply dissolved air flotation (DAF) where flotation is needed, and chemical coagulation using FeCl₃ or aluminium salts. This may evolve as efforts intensify to recover phosphorus from wastewater streams [211]. Additional electrochemical approaches include electro-Fenton processes (using iron-based catalysts) and hybrid systems combining, for example, EO with UV irradiation and/or catalysts such as aluminium oxide or graphene oxide; however, these approaches remain predominantly at the research and pilot scale.

3.7. Biocatalytic Treatment

A specialized form of biological treatment involves enzymatic reactors. A growing body of literature suggests that enzyme-based processes could provide rapid, potentially cost-effective transformation of poorly biodegradable trace organic contaminants at the tertiary stage of municipal wastewater treatment. A simplified process concept is shown in Figure 4. In this configuration, the primary role of the ultrafiltration unit is not direct removal of contaminants, but enzyme retention and recirculation to the bioreactor, since enzyme replacement would otherwise dominate operating costs.

As an alternative to ultrafiltration-based enzyme recovery, enzymes can be immobilised on support materials (e.g., polyester textiles) [212]. While immobilisation can simplify enzyme retention, it introduces additional challenges, including reduced catalytic activity and potential mass-transfer limitations. Typical trace organic contaminants (substrates), such as bisphenol A or F, sulfamethoxazole, diclofenac, and flufenamic acid, can be transformed by broad-specificity enzymes such as laccase C from Trametes versicolor. Because municipal wastewater contains complex mixtures rather than single compounds, studies that examine multi-substrate systems are particularly informative (e.g., [213]). Importantly, “remediation” in this context generally refers to transformation rather than complete mineralisation to CO₂: enzyme-mediated reactions can generate “different low- and high-molecular weight products” that are “difficult to analyse” [213] and may, in some cases, exhibit higher toxicity than the parent compounds. Enzyme activity is also sensitive to operational conditions, including pH, temperature, redox conditions, and the availability of mediators or co-factors.

Overall, enzymatic treatment is likely to be most suitable at point sources of trace organic contaminants (e.g., industrial effluents or hospital wastewater), where influent composition, temperature, and pH can be controlled and optimised for the intended enzyme–substrate system. At present, there is limited evidence supporting enzyme reactors as a robust final polishing step for municipal WWTPs.

The comparison presented in

Table 7. Comparison of treatment technologies for domestic wastewater pollutants.

Technology	Target Pollutants	Removal Efficiency	Advantages	Limitations
MF/UF	Microplastics (>1 µm)	High	Low cost, simple	Ineffective for dissolved pollutants
NF	Antibiotics, PFAS (long-chain)	Moderate–high	Lower energy than RO	Limited for monovalent ions
RO	PFAS, salts, pharmaceuticals	Very high	Broad removal spectrum	High energy, concentrate handling
MBR	Microplastics, organics	Very high	Combined bio + separation	Fouling, energy demand
Adsorption	Pharmaceuticals, PFAS	High (lab scale)	High selectivity	Limited scale-up
AOPs	Persistent organics	High	Degradation (not separation)	Energy/chemical demand
Electrochemical	Organics, metals	Moderate–high	No chemicals	Limited large-scale use

Abbreviations: MF, microfiltration; UF, ultrafiltration; NF, nanofiltration; PFAS, per- and polyfluoroalkyl substances; RO, reverse osmosis; MBR, membrane bioreactor; AOPs, advanced oxidation processes. ‘Electrochemical’ refers here to electrochemical treatment methods such as electro-oxidation and electrocoagulation.

highlights that no single treatment technology is universally effective for all pollutant groups. Low-pressure membrane processes are suitable for particle removal, particularly of microplastics, but are ineffective for dissolved contaminants. In contrast, high-pressure membrane processes such as reverse osmosis provide broad-spectrum removal, including salts and micropollutants, albeit at the expense of higher energy demand. Adsorption and advanced oxidation processes offer promising solutions for persistent organic contaminants, although their large-scale implementation remains limited. Therefore, the selection of appropriate treatment strategies should be based on the specific pollutant composition, the required removal efficiency, and economic constraints, often favouring hybrid or multi-stage treatment systems.

4. Current Outlook and Future Perspective

Several studies have already demonstrated the detrimental environmental effects of domestic pollutants, for example glyphosate derived from detergents and PFAS in toilet paper. However, additional monitoring studies of domestic pollutants are needed to support the development of efficient mitigation strategies, including regulatory measures and novel technologies.

Sodium chloride pollution represents a significant environmental burden. More effective, advanced, and salt-free softening technologies are already available [214,215]. Nevertheless, further research is needed to develop environmentally friendly domestic softening solutions and to understand the long-term water-quality changes associated with softened water, for example through AI-based water-quality monitoring [216].

In light of increasing scientific evidence on the environmental persistence and potential impacts of microplastics, the need for effective mitigation measures is growing. In this context, the REACH Regulation provides an important regulatory framework by pursuing a gradual phase-out of the intentional use of such microplastic particles. This approach is further embedded in broader policy initiatives such as the European Green Deal and the EU Strategy for Sustainable and Circular Textiles, which also address microplastic emissions from textiles. Consequently, the development of biodegradable alternatives, as well as product reformulation in sectors such as cosmetics and detergents, is expected to intensify, alongside an increasing need for scientifically robust analytical, detection, and standardised testing methods to assess microplastics and the biodegradability of polymers. In particular, requirements concerning the actual biodegradability of polymers under environmentally realistic conditions are becoming increasingly important, a development that is also reflected in the revision of the Detergents Regulation. In addition, further development of information and labelling requirements is expected to enhance transparency regarding product composition and environmental impacts [217,218,219,220]. These regulatory processes will also influence the development and adaptation of technological solutions.

On a global scale, phosphates/phosphonates are among the major domestic pollutant groups, as they are constituents of cleaning agents, detergents, and dishwashing products. While the environmental burden of phosphorus is classically associated with freshwater eutrophication, algal blooms, and oxygen depletion, current wastewater policy increasingly also addresses phosphorus as a recoverable resource. This policy shift is reflected at both EU and national levels [221,222]. Looking ahead, the strategic efficiency of technological solutions for phosphorus removal should be improved, including enhanced biological phosphorus removal (EBPR), chemical precipitation, and tertiary polishing where low effluent-P limits are required [223]. At the same time, efficient phosphorus recovery should be enabled.

A key issue for future discussion is that phosphorus recovery alone does not guarantee agricultural usefulness. For recycled phosphorus products, agronomic effectiveness and plant availability are critical evaluation criteria. For example, simple analytical indicators such as water solubility are not, on their own, sufficient to predict fertiliser performance reliably and therefore require refinement [224]. Phosphorus recovery from municipal wastewater should be assessed not only by recovery efficiency, but also by whether the recovered product delivers agronomically effective and plant-available phosphorus [225]. Among wastewater-derived recyclates, struvite appears particularly promising because it combines operational benefits, such as reduced scaling in wastewater treatment plants, with the production of a comparatively well-defined fertiliser product that has shown favourable plant-availability characteristics in both experimental and commercial-scale studies [225,226]. Finally, data-driven process intensification should be established for recovery processes; for example, complementary sludge- and sewage-sludge-ash-based recovery could be applied where broader plant-wide phosphorus recovery is targeted.

Domestic pollutants such as PFAS, pharmaceuticals, and PCPs, including other constituents such as heavy metals and microplastics, comprise a broad group of compounds of emerging concern. However, these groups remain poorly understood on a global scale, for example with respect to the accumulation of PFAS pollution in developing countries [227]. It is therefore essential to improve global monitoring of these compounds, for example through the introduction of novel sensor concepts [228]. As discussed in this review, novel technological concepts are also becoming available, but their economic performance still requires improvement, for example in the application of MOF–COF adsorption materials as more universal solutions for emerging contaminants. The practical feasibility of such technologies could also be enhanced through the development of novel sensor–AI frameworks [229].

5. Conclusions

First, domestic wastewater represents a highly heterogeneous mixture of pollutants, with large differences in both mass flows and environmental relevance among pollutant groups. In particular, high-volume inputs such as sodium chloride and household product use contribute significantly to overall pollutant loads, while low-concentration micropollutants, including pharmaceuticals and PFAS, remain critical because of their persistence and biological activity.

Second, the quantitative comparison based on per-capita emission estimates highlights that pollutant prioritisation cannot rely solely on concentration-based assessments. Mass-based approaches provide an important complementary perspective, enabling the identification of dominant emission pathways and supporting more targeted mitigation strategies.

Third, conventional wastewater treatment systems are not designed to remove many emerging contaminants effectively. While high removal efficiencies can be achieved for selected pollutant groups, persistent substances such as PFAS, antibiotics, and microplastics remain a major challenge, often requiring advanced or hybrid treatment technologies.

From a practical perspective, the findings underline the importance of combining technological solutions with source-control strategies. Reducing pollutant inputs at the household level, improving product design, and implementing targeted treatment technologies are all necessary to achieve meaningful reductions in environmental impact. In particular, high-mass pollutant streams may offer immediate mitigation potential, whereas persistent micropollutants require long-term technological and regulatory solutions.

Future research should focus on improving the quantitative understanding of pollutant flows across different regions, as well as on developing cost-effective and scalable treatment technologies capable of addressing complex pollutant mixtures. In addition, further work is needed to better understand the combined effects of multiple contaminants and to integrate these insights into risk assessment and regulatory frameworks. Addressing these challenges will be essential for the development of sustainable and resilient wastewater management systems.