Keywords

1 Introduction

Chemical pollution by pesticides, biocides, pharmaceuticals and personal care products, industrial chemicals, etc., poses a greater (cumulative) threat to environment. Personal care products are a varied group of compounds comprising preservatives (e.g., parabens), disinfectants (e.g., triclosan), fragrances (e.g., musks), UV filters/stabilizers (e.g., methylbenzylidene camphor, benzotriazoles), and insect repellants (e.g., DEET). Millions of consumers use cosmetic/personal care products and their ingredients on a daily basis to improve the quality of life. The unavoidable growth in the use of cosmetics/PCPs burdens the environment with their residues. The global production of personal care products is expected to reach 333 billion dollars by 2015 [1].

Although PCPs provide various benefits to the quality of life of the consumer, viz., soap, shower gels, toothpaste are to maintain hygiene and dental care, deodorants prevent body odor, and sunscreens protect human skin against adverse effects of UV light, they are generally excreted and emitted through the sewerage/wastewater system after use and ultimately released into nearby terrestrial or aquatic systems (Fig. 1).

Fig. 1
figure 1

Life cycle of PCPs in the environment with star (mark) showing the risk assessment (Adapted with modification from [2])

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Chemicals used in personal care products are biologically active compounds that are designed to interact with specific pathways and processes in humans and animals. A number of personal care products have been identified in environmental matrices and drinking waters [3–7], and their concentrations in environmental matrices are mostly in the range of ng–μg level. Many PCPs are environmentally persistent and bioactive and have the bioaccumulation potential. Thus, humans and terrestrial/aquatic ecosystems are greatly exposed to unknown cocktail of chemicals of parent as well as transformed products. Environmental (chemical) risk assessments of transformed products are rather complex than parent compounds due to scarce or nonavailability of toxicity data. The safety of a chemical in use is obviously based on a hypothetical zero-risk situation; however, that does not exist/or possible in a real-world situation. This peculiar, albeit unrealistic, aspect poses a major challenge for the risk assessment of chemicals and their ingredients/metabolites.

There have been a number of publications since the past few decades reporting on toxicity, fate, and transport of endocrine disrupting chemicals; nevertheless, information on residue levels and environmental risk assessment (ERA) of PCPs is scarce or nil until the end of the last century, and researchers started showing interest on analytical methods, bioaccumulation, and risk evaluation of PCPs only in the recent past.

Apart from the health risk to macroorganisms, the impact of PCPs on microbial community is still a question with few key outcomes. As we are aware, the prevalence of antibiotic-resistant bacteria in hospital, industrial [8], as well as domestic wastewater [9] environment is not uncommon; nevertheless, increasing use of antimicrobial compounds leads to similar problem of resistance in bacteria from sewage and surface water, drinking water, etc. [10–12]. Bacterial resistance for PCPs such as parabens in aquatic system is a growing environmental problem [11]. Moreover, a number of pollutants (i.e., pesticides, pharmaceuticals, illicit drugs, etc.) are continuously released into the environment, and their long-term effects on the receiving ecosystems are relatively unknown. Furthermore, interactions (synergistic/antagonistic) among the co-occurring compounds can also take place, complicating environmental assessment [13]. Considering the importance of PCPs’ emerging threat, this paper summarizes their risk assessment in the environment.

2 Pathways of Exposure and Uptake

The entry of PCPs into the aquatic environment includes direct disposal of domestic sewage and wastewater from hospitals and manufacturing industries, also they enter through wastewater treatment plant (WWTPs) effluents, leakage from septic tanks or leaching of landfill sites, and surface water runoff. The effluent and sludge from WWTPs and biosolids as manure shall be the prime source of PCPs in agriculture soil. The exposure of PCPs by organisms in the environment varies depending on the usage and resulting residual concentration/dilution in receiving waters, WWTP efficacy, and other possible exposure pathways.

The uptake of PCPs in aquatic ecosystem is mainly via contaminated water and secondarily by sediment. Some of the PCPs (e.g., triclosan) are ionizable substances, and the uptake of such ionizable substances depends on environmental conditions such as pH and soil/sediment characteristics. Mostly, the studies consider the bioavailability and uptake based on the properties such as octanol–water partition coefficient, bioconcentration/biomagnification factor, etc. [14]. However, no clear data on PCP uptake through food chain exists, so much research needs to be imparted to understand the real scale of PCPs bioavailability ([5, 14, 15] and references therein).

3 Methods of Risk Assessment

According to European commission [16], ERA is defined as an attempt to address the concern for the potential impact of individual substances on the environment by examining both exposures resulting from discharges and/or releases of chemicals and the effects of such emissions on the structure and function of the ecosystem.

Risk assessment identifies potential hazardous consequences of anthropogenic chemicals and determines the probability to occur in a specific environment (i.e., exposure assessment) and their severity (i.e., toxicity) [16]. Methods for assessing the ecological risks of anthropogenic pollutants are ample, and the most followed is the hazard quotient (HQ) approach [6, 16–18]. The quantitative approach to ERA includes three main components, viz., exposure assessment (predicted environmental concentration in different compartments such as water, soil/sediment, etc.), effect assessment (predicted no-effect concentration from dose–response relationship), and the risk characterization (calculating HQ). The hazard quotient or risk quotient (RQ) is calculated as the ratio between the predicted environmental concentration (PEC) or measured environmental concentration (MEC) and the predicted no-effect concentration (PNEC) in organisms [17]. The HQ/RQ values <0.1, 0.1–1, and >1 indicate low, medium, and high risks, respectively, of the individual compound [4].

The PEC for PCPs can be calculated based on multiple factors like type of substance, sales, population density, and usage statistics, and it may vary for each country and/or region. Nowadays, developed countries like the USA started using computational models (e.g., E-FAST) to predict the flux of PCPs in waterways [19]; nevertheless, it is quite difficult to calculate for developing countries where substantial statistics on production, sale, exact population, effluent load, etc., are hard to collect. In such condition, the relative MEC of specific compound is used instead of PEC. For calculating PNEC, most of the studies rely on either short-term acute toxicity (e.g., LC50, EC50, etc.) or long-term (sub-)chronic toxicity outcomes (e.g., no observed effective concentration (NOEC), lowest observed effective concentration (LOEC), etc.). Often, NOEC is calculated for individual organisms based on their toxicity endpoints; however, single NOEC representing multiple organisms (based on acute/chronic toxicity results) can be calculated by software such as ecological structure activity relationships (ECOSAR) of United States Environmental Protection Agency (USEPA). Indeed, for proper assessment, cumulative effect (chronic toxicity: growth rate, fecundity, abnormalities, etc.) is always preferred over one-time acute toxicity assay, because chronic data provides much better idea for the “true” risk of chemicals or chemical group and significantly lowers the use of uncertainty in risk assessment [20].

In risk calculation, an uncertainty/safety assessment factor (e.g., 10, 100, 1,000, etc.) is applied to acute or chronic toxicity endpoints to arrive at the PNEC. This application of uncertainty factor is based on the nature/form of toxicological data for different classes of organisms in each level of hierarchy/food chain. Usually, a safety factor of 1,000 is applied for acute toxicity endpoints, whereas safety factor of 10 is applied for chronic toxicity [17]. In general, among PNECs the lowest value for a specific taxonomic group was used to estimate the maximum risk posed by the chemical of concern [20].

The conventional PNEC calculated for a compound or stressor may not represent wider species assemblage or population (natural community). Therefore, to determine PNEC which is protective for most species/population/community, species sensitivity distribution (SSD) approach is followed, which represents the cumulative probability distributions of toxicity values from multiple species. Therefore, SSD is used in many instances [15, 21, 22], rather than conventional (single species) approach ([23] and references therein).

Jjemba [24] proposed an ecotoxicity potential (EP) to assess the extent of the risk of pharmaceutical and personal care products (PPCPs) based on fate (i.e., degradability), exposure factor (i.e., bioavailability), and effect factor (i.e., susceptibility) of the substance of concern.

$$ \mathrm{E}\mathrm{P}=T/V\left(\mathrm{NOEC}\right) $$

where T and V are the overall residence time and concentration of a substance in the environment, respectively. It is obvious that the lower the degradability (or the higher the persistence) and/or the higher the bioavailability of a chemical to nontarget organisms, the higher the magnitude of ecotoxicity potential.

Conventional HQ predicts risk based on MEC or PEC obtained from limited area and may not necessarily reflect a risk for larger ecosystem (e.g., entire river stretch). To fill the gap, environmental exposure models are developed to more precisely determine (weigh) the nominal exposure, for large area, over a period of time. Apart from PEC and MEC, exposure assessment models use variables such as the pathways of contaminant, form of the chemical(s) released, and its fate in different environmental compartments. Models like PhATE™ (Pharmaceutical Assessment and Transport Evaluation) and GREAT-ER (Geo-referenced Regional Exposure Assessment Tool for European Rivers) can be adopted for exposure assessments [25].

Apart from toxicity studies, computational approaches are gaining importance to replace/append the present risk prediction techniques (e.g., HQ), and one such approach is QSAR (quantitative structure activity relationship). Garcia et al. [26] performed the QSAR study using EPI SuiteTM interface, to understand the possible adverse effects of 96 PPCPs and metabolites with negligible experimental data and established a ranking of concern based on persistence (P), bioaccumulation (B), and toxicity (T) (extensive) of those PPCPs in Spanish aquatic environments. Their findings revealed that higher number of metabolites has got ranking equal to or greater than their parent compounds. Further, P, B, and T indexes are recommended recently by the Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) regulation to estimate the potential negative impact of chemicals on the environment [26].

Regarding PCPs, most of the studies either report the environmental concentration or its toxicological profile; however, only few studies were performed for risk assessment. In the present review, literature-based risk assessments of PCPs pertaining to HQ were primarily collected and grouped in Tables 1 and 2. The worst-case scenario reported for organisms in each of the study was taken for discussion. Further, the main purpose of the review was to collectively present the available ERAs of PCPs.

Table 1 Aquatic risk assessment of disinfectants and preservatives based on PNEC and MEC
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Table 2 Aquatic risk assessment of synthetic musks and UV filters based on PNEC and MEC
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4 Classification and Risk of PCPs

Regarding classification, each country adopts their own way of classification, e.g., sunscreens are cosmetics in the EU, whereas in the USA they are OTC drugs. Hair dyes are cosmetics in the EU but quasi-drugs in Japan, and their safety would be subjected to drug regulations necessitating drug-like safety dossiers [38]. Moreover, the PCPs can be grouped into categories based on their application (Fig. 2) such as antimicrobials (disinfecting agents and preservatives), insect repellants, fragrances (musks), UV filters/stabilizers, and siloxanes.

Fig. 2
figure 2

Major classes of PCPs with examples in parentheses and available ERA

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4.1 Disinfecting Agents

Disinfecting agents are antimicrobial compounds that are added as ingredients in sanitizers, disinfectants, and sterilants to control, prevent, or destroy harmful microorganisms (i.e., bacteria, viruses, or fungi). Since no single disinfectant is adequate for all situations, multiple disinfecting compounds are added in the formulations of PCPs [39].

Triclosan (TCS) (5-chloro-2-(2,4-dichlorophenoxy)phenol) and triclocarban (TCC) (3,4,4′-trichlorocarbanilide) are broadly used as antimicrobial and antifungal agents in household products of daily use (e.g., soaps, deodorants, skin creams, toothpaste and plastics, antimicrobial sprays, etc.). Due to extensive and inadvertent usage, residues of triclosan are ubiquitously found in surface water and sediment, WWTP influent/effluent, and fish ([6, 40] and references therein). Occasionally, fraction of TCS can occur as negative phenolate ion in environment due to its pKa (~8) and pH of the environment, which is considered to cause lesser toxicity than neutral (parent) form [41]. Further, Price et al. [41] opined that the risks of TCS calculated based on PEC/PNEC ratio will be an overestimate, so aquatic toxicity evaluation based on speciation is warranted.

Both TCC and TCS, having a log Kow of 4.2–4.76, are highly expected to get adsorbed onto solids and sediments and thus available for bioaccumulation [42–44]. Bioaccumulation studies showed that higher pH in environment can favor TCS bioaccumulation whereas lower pH could favor methyl-TCS to accumulate more [5].

Ecological risk assessment based on acute and (sub-)chronic toxicity tests was mostly available only for five antimicrobial agents in which TCS in river water from various countries (Switzerland, Japan, the USA, Slovenia, Spain, the UK, China, and India) showed high risk based on HQ for algae and most of the fishes and medium risk for crustacean (C. dubia) [6, 27]. Zhao et al. [28] reported high risk of TCS in Pearl River (Liuxi, Shijing, and Zhujiang rivers) water and sediment from China with maximum HQ observed as 23.4 and 28.7, respectively. Aside from rivers, Michigan lake and STP effluent in the USA were also found to contain the TCS at high risk level based on ECOSAR PNEC [7]. In addition to surface water samples, Kosma et al. [29] reported that TCS in WWTP effluents discharged into the rivers in Greece (Kalamas, Arachthos, Acheloos, Grevenitis, and Aliakmonas) may pose high risk to algae (HQ >100), fish (HQ >1), and invertebrates (HQ >1) in outfall locations.

Similar to TCS, TCC was also found at alarming level in river water (the USA and China), showing HQ >10 [27]. Zhao et al. [28] also indicated higher risk of TCC in water (HQ = 5.8) and sediment (HQ = 24.54) from the tributaries of Pearl River in China. In Michigan lake water (in the USA), medium risk was reported for TCC; however, effluent entering into the lake showed high risk (HQ >10) [7]. From Table 1, it is prominent that most of the HQs obtained for TCS (15 results out of 18) and TCC (6 results out of 7) were >1, pointing their risk in the aquatic environment is more likely. Among other disinfectants, resorcinol showed low (algae, P. subcapitata) and medium risk (C. dubia) for river water in Japan, whereas p-thymol and phenoxyethanol were found with low risk for daphnia, algae, and fish. This indicates that the risk from p-thymol and phenoxyethanol in Japanese rivers is minimal, unlike TCS and TCC [27].

Apart from risk assessment based on individual MEC, Reiss et al. [45] performed probabilistic exposure estimation based on transport and fate of TCS in wastewater effluents in the USA by using a model. The study compared the estimated exposure concentration with PNEC of most sensitive species of algae, plant, fish, and invertebrates and reported that some sensitive algae and plants may be at risk at effluent outfall with meager dilution. Further, the risk at downstream of the river is considered less because of dissipation of triclosan. While HQ is mostly derived from individual PNEC, some of the studies have generated common PNEC by SSD. Capdevielle et al. [15] constructed SSD based on chronic toxicity values for 14 aquatic species including fish, invertebrates, macrophytes, and algae and predicted lower risk of TCS to pelagic species immediately downstream of wastewater treatment plant discharge points in rivers of Europe (GREAT-ER model based on Calder river) and the USA (PhATETM model based on 11 catchment areas) by using a common PNEC of 1,550 ng/l. Further, Lyndall et al. [22] reported that 95th percentiles of measured and predicted TCS levels for water, sediment, and biota are consistently below the fifth percentile of the respective SSD, indicating no adverse effect of TCS.

The application of biosolids and wastewater containing TCC, TCS, and drugs to plant (soybean) showed higher accumulation of antimicrobials (at root tissue and beans) rather than drugs [46]; further it was reported that antimicrobials are not metabolized and thus accumulated whereas drugs can be eliminated/transformed by plants’ metabolism. So similar bioconcentration condition may favor the bioaccumulation of antimicrobials in aquatic food chain also. While there are ample reports on fate and risks of parental compounds, investigation on risk assessment of their derivatives/metabolites is scantily found. For instance, methyltriclosan, having greater hydrophobicity and bioaccumulation potential than triclosan, is less studied for its toxicity. Therefore, the environmental risk assessment may not be complete unless data on major derivatives/metabolites are also available.

4.2 Antimicrobial Preservatives

Among preservatives, parabens (alkyl esters of p-hydroxybenzoic acid) are widely used as bacteriostatic and fungistatic agents in cosmetic (creams, skin lotions, shampoos, soaps, toothpaste, etc.), pharmaceutical, and food industries [3, 31]. There are seven different types of parabens currently in use (benzyl, butyl, ethyl, isobutyl, isopropyl, methyl, and propyl). Although reports on environmental occurrence of parabens are ample ([3] and references therein), environmental risk assessment was scantily carried out [6, 30, 31].

Probabilistic risk assessment (PRA) of parabens in D. magna and fathead minnow was performed by Dobbins et al. [30] based on acute and chronic toxicity data. The observed HQs based on NOEC were much lower (7.8 × 10−6 − 2.3 × 10−4) than 1 (Table 1), which indicates no/little risk of parabens to fathead minnow and D. magna in surface waters of developed countries such as Belgium, Canada, and the UK [30]. Further, Yamamoto et al. [31] carried out an elaborate risk assessment for seven parabens in Tokushima and Osaka rivers in Japan. Unlike other studies, the NOEC values obtained from vitellogenin expression of fish were used, and the HQ showed no risk to aquatic organisms (algae, daphnia, and medaka) with the highest HQ obtained for n-propylparaben (0.01). Nevertheless, the sum of HQs of individual parabens showed low risk (HQ = 0.017) to those riverine organisms, and the PNEC based on n-butylparaben equivalence-based approach also showed low risk, with a maximum HQ of 0.018. They suggested that chronic tests at early life stages of fish are important for less erroneous risk assessment. Among developing countries, Ramaswamy et al. [6] evaluated the risk of four parabens in major rivers (Kaveri, Vellar, and Tamiraparani) of southern India. The lowest and highest HQs were observed for ethylparaben (8 × 10−6) and butylparaben (0.001) to fish, respectively. However, the calculated HQs for crustacean (D. magna) and fish (P. promelas) in all the rivers for all the parabens were below low risk criteria of 0.01.

4.3 Antioxidant Preservatives

Antioxidants are chemical substances used to prolong the shelf life of food items. Due to less stability of natural antioxidants, synthetic phenolic antioxidants (SPAs) like butylated hydroxytoluene (BHT) and butylated hydroxyanisole (BHA) are often preferred for their fat-soluble nature. The level of BHT was higher than triclosan and parabens in the rinse-off and leave-on cosmetics, respectively [47]. Their undisputed usage has resulted in trace quantities in food and environmental samples [26, 47]. Although BHA and BHT were classified as noncarcinogenic by the USEPA and safe food additives by the FDA and the EU, they possess estrogenic properties [48, 49]. Further, there are no environmental risk assessments available due to lack of toxicity data.

4.4 Insect Repellents

DEET (N,N-diethyl-meta-toluamide or N,N-diethyl-3-methylbenzamide), a broad-spectrum repellent and the most common active ingredient in insect repellents, is efficacious against mosquitoes and other insects of medical and veterinary importance. Till date, only few studies have reported acute toxicity in invertebrates, fish, and algae with EC50/LC50 in the range of 71.3–388 mg/l [5, 50]. Costanzo et al. [50] measured DEET residues in surface waters from Australia, Germany, the Netherlands, and the USA at safer level (75,000 times lower than EC50/LC50) for aquatic organisms such as algae (Chlorella prothecoides), water flea (D. magna), scud (Gammarus fasciatus), and fishes (Pimephales promelas, Gambusia affinis, Oncorhynchus mykiss). Further, Aronson et al. [19] estimated the flux of DEET in US rivers by iSTREEM (in-STREam Exposure Model) and E-FAST (Exposure and Fate Assessment Screening Tool) and predicted that DEET level was not expected to reach the lowest NOEC (521 mg/l) observed for algae, crustaceans, and fishes, indicating no risk of DEET in riverine habitat. Another insect repellent, 4-dichlorobenzene showed short-term exposure toxicity among invertebrates, fishes, and algae at lower concentration (1–60 mg/l) than DEET ([5] and references therein). Although newer repellents such as icaridin (1-piperidinecarboxylic acid 2-(2-hydroxyethyl) 1-methylpropyl ester) [51] and m-toluamide (N,N-diethyl-m-methylbenzamide) [52] are reported in the environment, their toxicity and risk assessment studies are not yet available.

4.5 Fragrances

Fragrances, the most widely used PCPs, seem to be omnipresent in the environment [3, 5]. Synthetic musks (SMs), being the most commonly used fragrances, are present in a wide range of products comprising deodorants, soaps, and detergents. Commonly used nitro musks are musk xylene (MX) and musk ketone (MK), whereas musk ambrette (MA), musk moskene (MM), and musk tibetene (MT) are used less frequently. In the case of polycyclic musks, celestolide (ABDI), galaxolide (HHCB), and toxalide (AHTN) are used most frequently, and traseolide (ATII), phantolide (AHMI), and cashmeran (DPMI) are used less often [3].

Although they are water-soluble compounds, due to high octanol–water partition coefficient of MK (log Kow = 3.8) and polycyclic musks (log Kow of 5.4–5.9), potential accumulation is expected in aquatic organisms. Rather than biomagnifications, direct impact on organisms is often understood by deriving HQ. In Nakdong River, South Korea, Lee et al. [32] reported low risk of toxalide and medium risk of galaxolide and musk ketone to fish. Combined risk of total SMs (Table 2) clearly indicates higher risk than individual, with higher contribution from MK. Apart from species-specific PNEC, Balk and Ford [33] used common PNEC to determine the risk of musks (AHTN and HHCB) in various environmental matrices, and the obtained HQ was always <1 (either no or low or medium risk). The risk characterized for AHTN based on NOEC for aquatic organism and fish-eating predators showed low risk (0.01–0.08), whereas medium risk (HQ = 0.44) was ascertained for sediment-dwelling organisms. For, HHCB, aquatic and sediment-bound organisms showed low risk, whereas no risk was determined for fish-eating predators. Interestingly, no risk was observed for worm-eating predators from HHCB (HQ = 0.001), even though medium risk (HQ = 0.1) was anticipated for soil organisms. Earlier, Tas et al. [53] also performed environmental risk assessment to understand the safety level of MK and MX in the Netherlands and found HQ ≤0.1 for both aquatic and sediment-dwelling organisms, while much lower HQ (0.01) was observed for fish-eating predators. Nevertheless, higher HQs were predicted for soil organisms with 0.5 for MK and 1.3 for MX, indicating medium to high risk, respectively and elevated HQ obtained for soil organisms implies the need for consideration of sludge being applied as fertilizer. Based on collective toxicity data and MECs, Brausch and Rand [5] suggested that probable risk for aquatic wildlife is more certain due to AHTN than other musks. However, chronic toxicity data on algae and benthic invertebrates are still lacking for effective aquatic risk assessment [5]. Apart from SMs, fragrances such as acetophenone, camphor, D-limonene, ethyl citrate, indole, isoborneol, isoquinolone, and skatole were also reported in surface waters. However, no acute/chronic toxicity data is available to evaluate their environmental risk [5].

4.6 UV Filters and Stabilizers

UV filters and stabilizers are found mainly in cosmetics and to some extent included in other personal care products, pharmaceuticals, food packaging, plastics, textiles, and vehicle maintenance products. Among organic and inorganic (ZnO, TiO2) variants, the organic forms are mainly used. Currently, 27 UV filters were designated for use in cosmetics, plastics, etc., and they are used in combinations (up to eight compounds) ([5] and references therein). The common feature of organic UV filters is the presence of an aromatic moiety with a side chain having different degrees of unsaturation and forming benzophenones (BPs), 4-methyl-benzylidine-camphor (4MBC), 3-benzylidine-camphor (3BC), homosalate (HMS), 2-ethyl-hexyl-4-trimethoxycinnamate (EHMC), ethyl-PABA (E-PABA), etc. After usage (showering, wash-off, laundering, automobile servicing, etc.), these chemicals enter the aquatic system indirectly (major input) from wastewater treatment plants and directly due to recreational activities such as bathing and swimming in lakes, rivers, and coastal waters (beaches).

In the environment, they may stay for longer duration because of high lipophilicity (log Kow 4–8) and poor biodegradability and eventually accumulate in sediments and biota as well [5, 54]. Like many xenobiotics, sunscreens do cause effects on aquatic animals [3, 55]. Danovaro et al. [56] reported that UV filters (commercial sunscreens, MBC, ethylhexylmethoxycinnamate, octocrylene, BP3, etc.) at very low concentrations cause rapid and complete bleaching of corals. The BCF for 4-MBC in roach, Rutilus rutilus, was calculated (9,700–23,000) ten times lower than methyltriclosan having similar log Kow (5) [55]. Due to potential bioaccumulation and toxicity, use of sunscreen products is now banned in some of the famous tourist destinations including marine ecoparks in Mexico ([56] and references therein).

Several studies have reported degradation of UV filters by photolysis ([57] and references therein), and the ecotoxicological data on parental compounds and their degradation products is scarce. Even though little information is available on their toxicity, environmental concentrations suggest low potential risk [58]. However, Gago-Ferrero et al. [58] presumed long-term risk associated with its pseudo-persistency in the environment. According to Diaz-Cruz and Barcelo [59], most of the UV filters and their metabolites are found to elicit hormonal (estrogenic and androgenic) activities based on bioassays (Fig. 3). Five compounds (including four BPs) showed high estrogenic activity, whereas only two showed high androgenic activity, and this indicates that UV filters possess endocrine disrupting potential.

Fig. 3
figure 3

Endocrine disrupting potentials of UV filters based on hormonal bioassays (no: activity not detected; low: submaximal dose–response curves with <30% efficacy; medium: submaximal dose–response curves with ≥30% efficacy; max: response curves with ≥80% efficacy) (From [59])

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Based on risk assessment of UV filters (Table 1), among benzophenones, BP1 and BP4 were found at levels to cause low risk (HQ >0.01) to fish and daphnia, respectively. Another benzophenone (BP3) showed medium, low, medium, ow risks to fishes, crustaceans, algae, and corals, respectively, indicating the variable toxicity expected in aquatic community. For the same organisms, 4-methyl-benzylidine-camphor showed low risk (HQ >0.01), except for algae with medium risk (HQ ≥0.1). Similarly, EHMC also pose low to high risks over a range of organisms; particularly, high risk was assumed for fishes. As reported by Fent et al. [35, 36], 3BC can cause serious risk to O. mykiss (HQ = 2.73) and D. magna (HQ = 1.43), and EHMC too pose a risk to D. magna (HQ = 1.35). Among other compounds, BP1 and E-PABA showed no risk, and BP2-4 showed low risk to aquatic species. Fent et al. [36] suggested the consideration of additive interaction of UV filters in mixtures for risk assessment; they investigated the acute (48 h) and chronic (21 day) toxicities on D. magna and found that acute toxicity increased with lipophilicity. In case of sea urchin (Paracentrotus lividus) and microalgae (Isochrysis galbana), medium (EHMC) to high risk (BP3 and 4MCB) was observed by Paredes et al. [37], and they opined that RQ is dependent on the selection of assessment factor which is still a debatable topic indeed.

Apart from the above compounds, 2-hydroxyphenyl derivatives of benzotriazoles (BTZs) are also one of the major groups of UV stabilizers reported in surface waters and biota ([60, 61] and references therein). Regarding the toxicological status, few studies are available based on acute studies which suggested BTZs and their derivatives are nontoxic with NOEC at few μg/l level for freshwater and marine organisms [60] and suggested for more chronic toxicity data for the organisms in different food chain for deriving any conclusion relevant to environmental risk assessments.

4.7 Siloxanes

Siloxanes used in many PCPs and industrial coatings are now ubiquitously reported in freshwater and marine environment [62–64]. Mostly, cyclic volatile methylsiloxanes (cVMSs), commonly called as cyclosiloxanes, are widely added as carrier solvents and emollients in cosmetics and other PCP formulations. Therefore, now the concern is about their potential toxicity, transport, and fate in the environment [65]. So far, cVMSs have received very little attention in ecotoxicological research, i.e., hazards and risks to aquatic biota. Wang et al. [63] reviewed the toxicological properties of octamethylcyclotetrasiloxane (D4), decamethylcyclopentasiloxane (D5), and dodecamethylcyclohexasiloxane (D6) with their respective log Kow (4.45, 5.20, and 5.86) and suggested strong tendency of cVMSs to bind organic matter in soil and sediment. Further, the BCF for D4, D5, and D6 were reported in the range of 1,875–10,000, 3,362–13,300, and 1,600, respectively, with most of the studies confirming the bioaccumulative (>2,000) and very bioaccumulative (>5,000) criterion ([63] and references therein). Further, Wang et al. [63] observed the most sensitive fish toxicity (acute/chronic) values for cVMSs in the range of 4.4–69 μg/l; however, to our knowledge, no ERA has been performed.

4.8 Antibacterial Resistance

Apart from the toxicity of antimicrobials to macro life forms, the more affected are the nontargeted microbes in the environment. It may hamper the bacterial diversity in environment, thereby affecting the community structure. Ricart et al. [66] demonstrated that environmental concentration of TCS can eliminate 85% of bacterial population at 500 μg/l level. Moreover, the biocidal effect [67] can trigger antibacterial resistance among the bacterial community. Evidences are growing on the prevalence of multidrug-resistant bacteria in the environment, drinking water, and patients, especially in developing countries such as India [68, 69], and the antibiotic-resistant genes (ARGs) have been isolated from the surface water, sewage, and in hospital environment [10]. Such conditions lead to the emergence/transmission of antibiotic resistance among bacteria in the environment [68]. Although the contribution of antibacterials in antibiotic resistance or multidrug resistance is largely unknown, the scientific committee on emerging and newly identified health risks by the EU [70] pointed that antibacterial resistance may develop rapidly when exposed to preservative(s). Therefore, uncontrolled and continuous use of antimicrobial/preservative compounds (triclosan, triclocarban, parabens, etc.) may lead to resistance in bacteria. Recent studies confirmed antibacterial resistance of PCPs (parabens, triclosan) from wastewater and surface water [11, 71]. Selvaraj et al. [11] reported bacterial resistance in common pathogens in effluents of sewage treatment plants in India for parabens and suggested the possible transfer of resistant genes to other pathogenic bacteria in natural waters because of the release of untreated wastewaters directly into the environment. Moreover, the resistant strains have the potency to modify PCPs into toxic compounds [72] which may further affect the organisms.

5 Present Risk and Future Prospective of PCPs in the Environment

On comparing the risk levels of major PCPs (Fig. 4), it is understood that most of the antimicrobials and UV filters showed medium to high risk whereas synthetic musks pose high risk only on total concentrations. Further, it clearly demonstrates that all the compounds within a group do not elicit similar toxicity but elicit cumulative risk. Apart from these three classes as shown in Fig. 4, reports on ERAs for antioxidants, fragrances, and siloxanes are lacking to be represented.

Fig. 4
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Aquatic health risk of PCPs based on literature data [6, 7, 27–37]

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In most of the studies, ERAs were performed based on individual compound and not for mixtures present in the environment; therefore, it is critical to assess and to understand their activity in mixture (combinations). Further, for more appropriate environmental risk assessment of PCPs, it appears essential to consider not only mixtures of parent compounds but also degradation products (metabolites, photodegradates, and chlorination by-products). This may pose an undefined ecotoxicological risk to resident organisms as well as a great challenge to ecotoxicologists. Moreover, testing chemical mixtures for toxicity is not an easy task due to the possible presence of thousands of organic and inorganic compounds (pollutants) in the environment; however, integrated dose–response relationships may be promising. In addition, the effect of individual components in the mixture can be extrapolated to understand/predict the cumulative effect via in silico approach, which can be further validated selectively through bioassays.

6 Conclusive Remarks

The impact of anthropogenic pollutants on the environment is severe and being given priority to understand well in this century. Despite their occurrence at submicrogram level in environment, the risk ascertained is quite high in many parts of the world. Developed countries are reporting high removal efficacy of WWTPs for few PCPs; however, higher risk is anticipated in developing countries where no proper treatment facilities are available. The exponential growth of population depletes freshwater resources and results in water shortage in this twenty-first century and in future. To combat water scarcity, reuse of wastewater is often advocated; such reuse has raised many questions with the occurrence of PCPs and other emerging chemicals residues.

Current environmental risk assessment procedures are limited in their proven ability to evaluate the combined effects of multiple xenoestrogens. Hence, ERA for mixtures (various forms of chemicals and their environmental derivatives) based on potential synergistic and/or antagonistic effects should be considered. As of the present situation, wider chronic toxicity studies should be imparted for many PCPs. Further, the effect of PCPs in the base of food chain may lead to adverse consequences through food chain magnification and ultimately on ecosystem. However, at present such scenario is entirely speculative and more appropriate studies to probe for this outcome have not yet been conducted holistically. Apart from the risk to aquatic organisms, some PCPs such as triclosan, parabens, etc., entering the aquatic environment may reduce the bacterial diversity and also act as buffers for the emergence of multidrug-resistant bacteria such as “superbug.” These concerns also need to be addressed for the safety of future generation.