Introduction

Landfill leachate is the liquid generated when rainwater percolating through landfilled waste combines with liquids produced during the decomposition of organic materials contained in the waste (Baucom 2013). It is a complex matrix characterised by a dark colour and strong smell, with high amounts of dissolved organic matter, macro inorganic compounds, heavy metals, and organic compounds (Kjeldsen et al. 2002; Baun et al. 2004; Ward et al. 2005). Reports by Wowkonowicz and Kijeńska (2017) and Qi et al. (2018) show that some leachate constituents, especially the organic compounds have endocrine disruptive capabilities implying they could interfere with hormone synthesis, and hormone mediated reactions, and responses in humans and animals. Compounds with these characteristics are referred to as endocrine disruptive compounds (EDCs).

EDCs could either be natural such as steroidal (progestogen, androgens, oestrogens) and non-steroidal hormones (phytoestrogens), or synthetic like the phthalates, polyhalogenated and phenolic compounds, pesticides, pharmaceuticals and personal care products (PPCPs) and their respective metabolite most of which have anthropogenic origins (Wee and Aris 2017). Exposure of humans and animals to EDCs according to Swedenborg et al. (2009) and the United States Environmental Protection Agency (USEPA 2016) could affect hormone synthesis, secretion, transport, binding action, and elimination in the body, with consequent negative health outcomes. Hence, research on the occurrence and characteristics of EDCs has gained traction in the last few decades.

This review article focuses on two known EDCs: phthalates and phenols. Phthalates (also referred to as phthalic acid esters (PAEs), and dialkyl or alkyl esters of 1,2 benzene carboxylic acid), are produced when alcohols react with the carboxyl functional group on the benzene ring of phthalic acids (Opeolu et al. 2010; Adeniyi et al. 2008). They are classified as either high molecular weight phthalates (8–13 carbons in the carbon chain) among which are di-(2-ethylhexyl) phthalate, Di-n-octylphthalate, diisononyl phthalate, and dibutyl phthalate, or low molecular weight phthalates (3–7 carbons in the carbon chain) which include compounds like dimethyl phthalate, diethyl phthalate, and di-n-butyl phthalate (Schettler 2006). Phenols on the other hand, are compounds containing a benzene ring unto which a hydroxyl (OH−), a methyl (CH3), an amide (NH2), or a sulphonic group has been bound (Slack et al. 2005; Qi et al. 2018). Examples of phenols include bisphenol A (BPA), cresol, 2,4 dichlorophenol, nonylphenol, and nitrophenols among others. Both phthalates and phenols are widely used in the manufacture of several products, especially plastic products that are needed by humans for various purposes in their everyday life and are therefore present in wastes that are disposed in landfills. Both groups of compounds are listed among priority pollutants due to their persistence in the environment and potential toxicity to humans especially with regards to endocrine disruption and carcinogenicity. Efforts to reduce their spread in the environment and to remediate environments where they are present are in overdrive.

Most research on EDCs have focused on wastewater from various sources. Lange et al. (2014) and Wee and Aris (2019) for example studied the occurrence and elimination of EDCs in municipal wastewater treatment plants. Gadupudi et al. (2021) and Azizi et al. (2021) reviewed current technologies used to remove EDCs from wastewaters. Kumawat et al. (2022) investigated seasonal disparities in the occurrence of EDCs in drinking water supply systems. Other studies on EDCs in wastewater can be found in Cesaro and Belgiorno (2016), Becker et al. (2017), Ojha et al. (2021), Rodríguez-Hernández et al. (2022), and Budeli et al. (2022) among others. Little attention has been given to EDCs in landfill environments, yet this is also a major source of these compounds, given that they are contained in most products used at homes, medical and academic laboratories and institutions, and industry. Some of these studies like those of Mahiudddin et al. (2012), Li et al. (2019), Boll et al. (2020), Chang et al. (2022), Asimakoula et al. (2023), Ateş and Argun (2023), Patil and Jena (2023) and Ye et al. (2023), have mostly been carried out in laboratory settings which are usually much simpler and less complex and dynamic relative to landfill and soil environments.

This article presents a review of the sources, occurrence, and fate of phthalates and phenols in landfilled wastes and landfill leachate, and the conditions under which phthalates and phenolic compounds migrate from waste into landfill leachate. The fate of phthalates and phenols in soils in the vicinity of landfills is also discussed in this article. An understanding of these aspects will contribute towards effective remediation of environments contaminated with these priority compounds. The article concludes with a discussion on the health implications of human exposure to phthalates and phenolic compounds, issues to be considered in their remediation, and identifies information gaps that are needed to gain better understanding of the behaviour of phthalates and phenols in landfills and surrounding soil environments with a view of developing more integrated management strategies to curb their spread.

Methods Used to Source Information

This review made use of a narrative and scoping approach to gather information relevant for the review and the identified problem. According to Munn et al. (2018), scoping is relevant when information on a selected topic is still hazy. Through scoping, information on phenols and phthalate occurrence in landfills and surrounding areas were revealed. Their fate in wastes, landfill leachate and surrounding soils were discussed, and the characteristics that influence the fate of these compounds in the environment were highlighted. Articles consulted for information spanned mostly between 1990 and 2024, with a few articles found to be relevant from previous years included. Journals were consulted without paying particular attention to a specific database and so information from diverse databases were included. The key words used for sourcing information are included in Table 1. Whereas the focus was on phthalate and phenolic compounds generally, the search considered individual compounds belonging to these chemical groups. Information obtained through these searches was then reviewed by the authors for relevance prior to their inclusion in the review.

Table 1 Search criteria used to identify relevant articles to be consulted
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Phthalate and phenol migration from landfilled waste

According to Giuliani et al. (2020), and Kotowska et al. (2020), phthalates are contained in plastic packaging, nail polish, shampoos, cosmetics, after shave, detergents, surfactants, wood preservatives, herbicides, plastics, adhesives, batteries, alloys, ceramics, and electrical appliances among others. Phenols on the other hand are used in the manufacture of pesticides, insecticides, dyes, explosives, wood preservatives, antioxidants, adhesives, herbicides, dispersants, detergents, emulsifiers, pulp bleaching, and as anti-foam products (Slack et al. 2005; Boonnorat et al. 2014). Kurata et al. (2008) have also highlighted the use of phenols in the manufacture of nylon, chemical skin-peelers, nerve injections, topical anaesthetics, and disinfectant, and they have also been detected in antiseptics, lozenges, lotions, ointments, paints, polishes, adhesives, lacquers, varnishes, and solvents. Most of these products are found in household wastes and are therefore likely to find their way into landfills. Once in the landfill, these compounds could leach out of the waste materials containing them because of various chemical and biochemical processes.

Knowledge on how these compounds get leached out of materials containing them in the landfill is scarce most probably because the landfill is a closed environment where wastes are compacted at the end of each day, making it difficult to monitor any chemical or biological process going on in the landfill in real time. In addition, landfills are characterised by huge variations in the waste types received and the prevailing chemical and biological processes going on, making broad generalisations unreliable. However, the fact that phthalates and phenols have been reported in landfill leachates in many studies indicates that they are being released from the waste materials buried in the landfill. Most studies on phthalates and phenols around landfills environments have looked at their occurrence and degradation in leachates albeit mostly under simulated conditions. Considering that few studies have been carried out on the leaching of phthalates and phenol from waste materials buried in landfills, insights on the leaching of contaminants from buried wastes could be inferred from the outcome of studies which have focused on the migration of these contaminants and their congeners from materials containing them. Examples of such studies include the migration of phthalates and phenols from packaging material to various food substances. Another source of insight could be from information on bonding between these compounds and materials containing them or how they are incorporated in these materials.

The migration of phthalates from non-food plastic packaging and plastic food packing to food have been reported by Same et al. (2023) and Yang et al. (2019) respectively. Bisphenol A migration from cans with epoxy resin linings to water was also reported by Fasano et al. (2012). The migration of alkylphenols from packaging to fruit juices was reported by Shabani et al. (2023). Eckardt et al. (2020) showed the release and migration of phenolic compounds in polyester-phenol coatings used for packaging. Their results all showed that these compounds migrate from the materials containing them under various circumstances and the extent of migration varied with time and temperature. Most of these studies used solid phase extraction to extract the samples from the food materials and the percentage recovery of the compounds from the material investigated ranged from 75 to 110% with limits of detection and limits of quantification of 0.4 μg/l and 0.9 μg/l respectively.

The migration of phthalates like diethylhexyl phthalate (DEHP), dibutyl phthalate (DBP), diethyl phthalate (DEP) and butyl benzyl phthalate (BBP) from plastic was shown to be initially slow for the first 50 days but increased exponentially thereafter (Same et al. 2023; Ayamba et al. 2020). Heat and changes in pH conditions were reported by Fasano et al. (2012) to increase the hydrolysis of the ester bonds linking BPA molecules in epoxy resins facilitating their migration into foods. Further reports on the migration of phthalates and phenols from plastics can be found in Bosnir et al. (2007), Fasano et al. (2012), Andrade et al. (2019), Liu et al. (2020), Zhang et al. (2022), Geueke et al. (2023), Wang et al. (2023), and Đurić et al. (2024). According to Adeniyi et al. (2008), Liu et al. (2010), Domínguez-Romero et al. (2023) and Wang et al. (2023), phthalates are not chemically bonded in the materials where they are found and so they can be easily leached out, or washed off from products containing them. Phenols like BPA are easily leached out from plastics where they are covalently bonded to other compounds in the plastics compared to plastics in which they are added to the compounds by mixing (Manchester-Neesvig et al. 2001). From these studies, one can infer that phthalates and phenols are likely to leach from buried waste materials in landfills at rates which vary with several factors.

Factors Influencing Phthalates and Phenol Migration from Buried Wastes

Leachate Characteristics

Studies by Bosnir et al. (2007) looking at the migration of phthalates from plastic containers to soft drinks have shown that acidic pH conditions facilitates the migration of phthalates from plastics but Teuten et al. (2009) have indicated that when an acidic leachate with high ionic strength surrounds landfilled waste, the potential of the waste materials to release organic compounds is lower than when the leachate is neutral or alkaline. This inconsistency could be attributed to the fact that landfill leachate and most other natural environments containing materials rich in these compounds are more complex than soft drinks because of the presence of other chemical contaminants such as inorganic salts and organic solvent which may interfere with the interaction between the leachate and the waste materials in the case of a landfill. For example, Xu et al. (2011) found enhanced leaching of BPA from plastics and cardboard with increased dissolved organic matter (DOM) under neutral conditions and this has been attributed to the strong affinity of phthalates for the humic fraction of the DOM (Zheng et al. 2007). High salinity on the other hand reduces the leaching of phthalates from microplastics. Considering that dissolved salt content, and hence salinity levels in landfill leachate is usually high, it could interfere with the release of phthalates from waste under acidic conditions. These findings further highlight the complexity of the landfill environment and landfill leachate and indicate that there are other components in alkaline and neutral leachates that enhance their ability to extract phthalates and phenols from waste materials.

Since alkaline leachates are usually present in older landfills, and the migration of phthalates increases with time (Same et al. 2023), the leaching of phthalates from waste in the landfill is therefore expected to increase as the landfill ages. This was confirmed by Sajiki and Yonekubo (2003), Kurata et al. (2008), and Teuten et al. (2009) who all reported that methanogenic leachates tend to extract more phthalates and phenols from waste materials than acidic leachate, and Masoner et al. (2014), who reported higher concentrations of phthalates in closed compared to active landfills. Similar observations were made by Yazici et al. (2011) for phenols. Studies on the effect of pH on the leachability of phenols from waste have been mostly on BPA where its concentrations in landfill leachate increased with increase in pH. This was not the case with the cresols, which suggest that the effect of pH on the concentration of phenolic compounds in landfill leachate is mostly observed in phenols with lower molecular weight, which highlights the role of the structure of the compound involved in its leachability from landfilled waste.

Compound Properties

Phthalates are generally hydrophobic, a property which may reduce their solubility from waste contained in the landfill, but their solubility varies with the specific phthalate. Teuten et al. (2009) showed that dimethyl phthalate (DMP) is easily released from waste products where it is found because of its high-water solubility. DEP and DMP are weakly hydrophobic compared to DEHP (Bauer and Herrmann 1998) as shown by their octanol/water partition coefficient (logKOW) (Table 2) and Asakura et al. (2004) noted that DMP and DEP are easily leached out from waste than the more hydrophobic DEHP. One would therefore expect higher concentrations of DEP and DMP in landfill leachate than DEHP but that is usually not the case as the data in Table 2 shows. Several reasons could explain this observation. According to Lee et al. (2020) and Liu et al. (2010), DEHP is strongly absorbed onto organic matter, hence it could easily accumulate in landfill leachate which is rich in organic matter. Further to this, the half-lives of DMP, DEP, di-n-butyl phthalate (DnBP) and DEHP in natural waters under aerobic biodegradation are 0.2–10, 0.3–12, 1.7–2.5 and 2–22 days respectively (Cousins et al. 2003), which indicates that DEHP is more persistent in aquatic environments than other phthalates. In addition, DEHP is the most widely used phthalates and so its concentration in the environment is likely to be higher than that of other phthalates.

Table 2 Sources and physical properties of some phthalates and phenols identified in landfill leachates
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Data presented in Table 2 indicate that phenols are relatively more soluble in water than phthalates as reflected by their higher values for water solubility. The LogKOW values of phenolic compounds are also relatively lower when compared with those of the phthalates and as such, a higher concentration of phenols is expected in leachates than phthalates because of their weak hydrophobicity. Bisphenol A which has the highest concentration in leachates among the phenols according to data in Table 2 has a logKOW of between 3.32 and 3.82 making it moderately soluble in water. It is also relatively less hydrophobic compared to other phenols and so it can easily migrate from waste which could justify its high concentrations in landfill leachates (Tables 2 and 3).

Table 3 Descriptive statistics of concentrations (µg/l) of commonly encountered phthalates and phenols in landfill leachates
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The structure of the compound also plays a significant role. Teuten et al. (2009) for example have shown that phenols with shorter alkyl chains are more easily leached from waste materials than those with longer chains. Waste materials containing products rich in BPA, cresols, and dimethyl phenols are therefore likely to contribute a higher percentage of phenols in landfill leachates than those containing compounds like octyl, nonyl- and butyl phenols because of the differences in the length of their alkyl chain. The frequent use of these compounds, their characteristics, and structure all work together to determine which compounds are likely to occur in landfill leachate.

Environmental Factors

The prevailing environment surrounding waste in landfills including the presence of suitable solvents, pH, and temperature, salinity, pressure, and dissolved oxygen influence the release of phthalates and phenols from wastes in landfills (Zhao et al. 2018; Same et al. 2023; Henkel et al. 2023; Benjamin et al. 2015; Hussain et al. 2011; Nair et al. 2008; Xu et al. 2011; Wojnowska-Baryła et al. 2022). Studies by Ayamba et al. (2020) show an increase in migration of DEHP and DBP from polythene plastic materials with increase in temperature, with the amount of DEHP migrating from the plastic doubling between 5 and 80 ℃. This could be of significance in the migration of phthalates in wastes buried in landfills where temperatures could increase to these levels because of microbial activities and various degradation processes. The presence of organic solvents in the leachates could also enhance leaching of these organic contaminants from waste. Phenols and phthalates identified in landfill leachate are therefore the results of a complex series of interactions between waste, leachates, microbial activities, and prevailing environmental conditions which facilitate their migration from waste materials buried in the landfill. Given that no two landfills have the same conditions, only broad generalisations about the migration of these contaminants from buried wastes can be made. Further investigations are therefore required to better understand how, and the ease with which these compounds can leach out of landfilled waste materials into leachate.

Occurrence of Phthalates and Phenols in Landfill Leachate

Phthalate compounds and their metabolites that have been identified in landfill leachates include butyl benzyl phthalate (BBP), mono benzyl phthalate (MBzP), di-n-butyl phthalate (DnBP), diethyl phthalate (DEP), and di-n-octyl phthalate (DnOP), di-(2-ethylhexyl) phthalate (DEHP), di-isodecyl phthalate (DiDP), di-isononyll phthalate (DiNP), mono-(2-ethylhexyl) phthalate (MEHP), monoethyl phthalate (MEP), di-n-hexyl phthalate (DnHP), mono-n-butyl phthalate (MBP), and mono-isobutyl phthalate (MiBP) (Bauer and Herrmann 1998; Wowkonowicz and Kijeńska 2017; Al Raisi 2022; Pisharody et al. 2022). This list, which is inexhaustive, indicates that both the monoesters and diesters of phthalates are present in landfill leachates. However, the phase in which the landfill is determines whether the monoesters or the diesters will be dominant. In the degradation of phthalates, the diesters are usually degraded into the monoesters by micoroganisms and so one would expect high concentrations of monoesters in leachates in landfills that are in the methanogenic phase when microorganisms are actively breaking down the diesters. The phase of the landfill could therefore influence the types of phthalates that could be present in its leachate.

Among the phthalates, DEP, DBP, and DEHP seem to be the most frequently encountered in landfill leachates whereas DMP is less so. In studies conducted by Kotowska et al. (2020) on landfill leachates in Poland, DEHP was the most encountered phthalate occurring in majority (97%) of the leachate samples characterised whereas DIBP, DIHP, DEP, and DBP only occurred in in 58–81%, and DMP, DPP, DINP in less than 42% of samples. A review of phthalates in leachates from landfills in different countries by Wowkonowicz et al. (2021) also showed that DEHP was present in leachates from 5, 14, 3, 1, 2, 1, 1, 1, 11, 2, 5, and 6 different landfills in China, Sweden, Poland, Canada, Japan, Puerto Rico, Thailand, Finland, Italy, Germany, and Denmark, respectively, and at different levels which highlights the prevalence of DEHP among the phthalates in landfill leachates. These observations were confirmed by analyses of phthalate concentrations reported in landfill leachates from different countries presented in Table 3 where DEHP was encountered in up to 51 landfill leachates studied whereas DEP and DBP were encountered in 31 and 29 respectively.

DEHP does not only occur frequently but it also has the highest concentration among the phthalates commonly encountered in landfill leachate as shown by the data analyses in Table 3. Gholaminejad et al. (2024) reported a concentration pattern of DEHP > DnBP > DnOP > DEP > DMP in landfill leachate from Iran whereas in another study on various landfills in Europe by Jonsson et al. (2003a), phthalic acid, DEHP, BBP, BDP, DEP, monomethyl phthalate (MMP), monoethyl phtshalate (MEP) and Monobutyl phthalate (MbutP) were the commonly encountered phthalate compounds. No DMP was identified in these landfills. Studies by (Kalmykova et al. 2013) reported ranges of < 1–5 μg/l, < 1–4 μg/l, < 1 for DEP, DnBP, and BBP respectively in leachate from four different landfills in Sweden. DEP and DEHP with concentrations of between 2 and 400 μg/l, were also reported in landfills in the USA by (Masoner et al. 2014) and they also noted that these compounds occur more frequently (80%) in leachates from closed landfills compared to active landfills (59%). (Bauer and Herrmann 1997) in their studies on the leachability of PAES from domestic waste found that DEHP contributed up to 91% of the sum of phthalates determined in leachates whereas DBP made up 8%, with DMP, DEP, and BBP making up the remaining 1%. These studies highlight a high variability in the concentrations of phthalates in landfill leachates which is confirmed by the values of variability of leachate concentrations in Table 3 but also show that a few phthalates (DEP, DBP, and DEHP) are consistently occurring in landfill leachates. High concentrations and prevalence of DEP, DBP, and DEHP in landfill leachates could be explained by the fact that these compounds are widely used for the manufacture of everyday goods and the waste materials containing them are solids whereas those containing DMP which occurs less frequently are usually liquid products which are disposed along with wastewater.

The occurrence of phenols in landfill leachates is less common compared with the phthalates and even when they occur, their concentrations are much lower than what is reported for phthalates. This is reflected in Table 3 where the landfill leachates in which phenols have been identified are fewer than those in which phthalates have been identified. Most of the products reported to contain phenols are likely to end up in the wastewater stream because they are mostly liquids which may explain the lower concentrations of phenols in landfill leachate compared to phthalates which is a main component of many solid products, especially plastic materials disposed in landfills. The containers of these liquid products are however usually disposed together with solid waste in landfills and may also be contributing to the load of phenols in landfill leachates. Another source of phenols in landfill is the breakdown of other polyaromatic compounds and lignin that may be present in landfilled waste, which produce short chain non-halogenated alkyl- and carboxylated phenols (Eggen et al. 2010). Leachates from landfills are therefore expected to contain phenols and their derivatives.

Some of the phenols commonly encountered in leachate from landfills in different stages of decomposition include 4-nonylphenol, bisphenol A, 2,4-dimethylphenol; 4-tert-butylphenol; 4-tert-octylphenol; 2,4-/2,5-/2.6-dichlorophenols; 2,6-dichlorophenol; 2,4,6-trichlorophenol; 2,3- 4,6-tetrachlorophenol; pentachlorophenol, and cresols (Yazici et al. 2011; Slack et al. 2005). Jaradat et al. (2022) reported phenol concentrations of up to 8.8 mg/l in landfills in Jordan. Kalmykova et al. (2013) reported ranges of 0.0–5.7 μg/l for 4-nonylphenol, 0.0–2.4 μg/l for nonylphenol and 0.0–15 μg/l for octyl phenol in landfill leachates in Sweden. In studies carried out by Reitzel and Ledin (2002) in Danish landfills, the concentrations of phenols identified followed the order o/p cresol (29 μg/l) > 3,5-dimethyl phenol (27 μg/l) > m-cresol (17 μg/l) > 2,4-dimethylphenol (13 μg/l) 3,4-dimethylphenol (10 μg/l) > 4-chloro-o/m-cresol (4.6 μg/l) > 2/3-chlorophenol (1.6 μg/l) > 4 chlorophenol (1.3 μg/l). Trichlorophenol was not detected in most of the landfills studied. Among the phenolic compounds frequently encountered in landfill leachates, the cresols, 2,4-/2,5-dichlorophenols, phenols, 4-tert-octylphenol, 2,6-dichlorophenol, 4-tert-butylphenol, 2,4,6-trichlorophenol, pentachlorophenol, 2,3,4,6-tetrachlorophenol and 2,4-dimethylphenol have been listed among those with high concentrations as the studies by Kurata et al. (2008) and (Reitzel and Ledin 2002) showed with cresol, BPA, phenol and octyl phenol having the highest concentrations (Tables 2 and 3). An analyses of phenols in landfill leachates around the globe shows a wider range in concentration than the phthalates and except for BPA, the variability is also lower (Table 3). These analyses show that BPA and phenol are the most frequently reported in landfill leachates. Though cresol is less commonly encountered, when it occurs, its concentration is usually higher than that of phenol but lower than that of BPA (Table 3).

Data reported in various studies including those in Table 2 show that the concentrations of some phenols and phthalates in landfill leachate exceed the maximum acceptable concentrations permitted in water meant for drinking according to WHO standards. Examples of these compounds include DBP, DEP, DEHP, and cresol whereas others like nonylphenol and dichlorophenol are usually below the acceptable limits. It must however be noted that most of these compounds do not have threshold limits, but they have all been included on the list of priority compounds in drinking water because of their potential human and environmental impacts and risks. Their detection in landfill leachate could also be related to the method used for their extraction.

Methods Used in Analyses of Phthalates and Phenols in Landfill Leachates

Both solid phase extraction (SPE) and liquid–liquid extraction (LLE) have been used to extract phthalates from landfill leachates using a variety of solvents with varying recovery efficiencies. In studies that have used LLE, solvents like methanol, dichloromethane, hexane, acetonitrile among others are commonly used to extract the phthalates and phenols from leachate whereas in those studies where SPE has been used, various columns packed with different absorbents including silica gel, alumina and graphatized carbon among others have been used to remove the compounds from the leachate and later desorbed using various solvents. Though the columns used for SPE and the solvents used for LLE in these analyses vary, good percentage recoveries have been reported in both cases.

Beldean-Galea et al. (2013) for example showed good recovery of phthalates (105%) from landfill leachate when LLE was used with dichloromethane as solvent whereas using SPE, Olujimi et al. (2010) reported recoveries of between 97 – 101% for phthalates. Wowkonowicz and Kijeńska (2017) reported 35% uncertainty with the determination of phthalates using LLE with solvents like acetone/dichloromethane, followed by n-hexane. Percentage recovery of these compounds from landfill leachate however seems to be compound specific. The recoveries of phthalates in studies carried out by Kotowska et al. (2020) using SPE followed the order DPP > DEHP > DEP > DBP > DiBP > DMP with recovery percentages of 137, 132, 119, 115. 76 and 15% respectively whereas in another study on three phthalates by Sun et al. (2013) and Boonnorat et al. (2014) also using SPE, the order of percentage recovery was DMP (105%) > DEHP (97%) > DEP (93%). Whether using SPE or LLE, DEHP shows a high percentage recovery and limit of quantification than the other phthalates. Studies on the extraction of phthalates from various environmental samples can be found in Wowkonowicz and Kijeńska (2017), Sun et al. (2013), Boonnorat et al. (2014) and Kotowska et al. 2020). Phenol extraction from landfill leachate has also been done using both SPE and LLE. Percent recoveries were < 50% for phenol and 109% for BPA when dichloromethane was used as solvent (Beldean-Galea et al. 2013). Using SPE with acetonitrile and methanol as solvent, percent recoveries of between 26% for phenol and 104% for trichlorophenol were observed by Reitzel and Ledin (2002).

Though LLE and SPE are commonly used, their efficiencies diminish in situations where the concentrations of the compounds are low and so the samples need to be enriched with the analyte of interest prior to extraction to maximize recovery. This step is being eliminated by the use of solid phase microextraction (SPME) which is an adsorption or absorption based solvent-free technology that makes use of coated fibre or Arrow to concentrate compounds in a sample which could be volatile or semi volatile (Weggler et al. 2020). In the analyses of phthalates and phenols, SPME is used to preconcentrate and extract the compounds of interest in the same step with no solvents required. Studies that have used SPME to determine phthalates in various environmental samples (Alshehri et al. 2022; Polo et al. 2005) have reported percentage recoveries of between 72 and 109% with the fibre type used for extraction playing a minimal role in the level of recovery.

The extracted phthalates and phenols are commonly analysed with Gas chromatography mass spectrometry (GC–MS) which has its detection limits for these compounds in the lower μg/l level (Olujimi et al. 2010). There is therefore good percentage recovery of phthalates and phenols from landfill leachate and so the patterns of concentrations of the different phthalates and phenols reported in different studies cannot be blamed on the methods but seem to be a true representation of what exists in most landfill leachates. Phenols concentrations in landfill leachates are therefore generally lower than those of phthalates. Studies where the same leachate sample is extracted using LLE, SPE, SPME, and microwave are needed to determine any differences that may exists because of the extraction methods.

Fate of Phthalates and Phenols in Landfill Leachate and Influencing Factors

Phthalates

When waste is disposed in a landfill, decomposition is first aerobic for a short while after which, anaerobic conditions assume prominence as the oxygen buried with the waste is rapidly depleted (Slack et al. 2005). Under anaerobic conditions, hydrolysis and fermentation processes break down organic compounds in the leachate into their metabolites, carboxylic acids, and alcohols (Slack et al. 2005) depending on the properties of the compounds. Biodegradation of phthalates occurs mainly in the upper layers of the landfill possibly due to the presence of microorganisms at these layers, whereas in the lower layers, hydrolysis dominates (Huang et al. 2013). During biodegradation, phthalate diesters are converted to paracatechuate as summarized in Fig. 1 (Jianlong et al. 2000; Jonsson et al. 2003a; Schwarzbauer et al. 2006; Gao and Wen 2016). The paracatechuate is then degraded through either the meta- or ortho cleavage pathway (Boll et al. 2020).

Fig. 1
figure 1

Fate of phthalates under both aerobic and anaerobic conditions

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Bacteria involved in aerobic degradation of phthalates are mostly the rod-shaped bacteria belonging to the Roteobacteria, Actinobacteria, Firmicutes, Chlorobi and Deinococcus-Thermus divisions (Kapanen et al. 2007; Çevik et al. 2019), and include organisms such as Pseudomonas sp, Burkholderia sp, Arthrobacter sp, Sphingomona sp, Ochrobacterun sp, and Acinetobacter sp, most of which are facultative aerobes (Gao and Wen 2016). Among the fungi are Aspergillus parasiticus, Fusarium subglutinans, and Penicillium funiculosum (Pradeep and Benjamin 2012). Anaerobic phthalate degrading bacteria include Clostridium sp, Bacillus sp, Pelotomaculum sp, and Pseudomonas species (Gao and Wen 2016). The degradation of phthalates is however slower when conditions become anaerobic as only the side chain (-CH3) of the phthalate is broken off, while the aromatic ring remains intact (Nozawa and Maruyama 1988).

At the bottom of the landfill, hydrolysis occurs in two steps whereby, the phthalate diester is first converted to a monoester and a corresponding alcohol moiety, and then to phthalic acid and a second alcohol (Huang et al. 2013). Hydrolysis in the landfill is strongly correlated with the methanogenic flora in the landfill (Jonsson et al. 2003b), and is facilitated by enzymes secreted by these microbes which break off the side chain of the phthalates (Bauer and Herrmann 1998). However, abiotic hydrolysis is the more likely process of phthalate breakdown taking place in the landfill as opposed to biotic hydrolysis according to Bauer and Herrmann (1998). The main factors driving abiotic hydrolysis of phthalates in the landfill are the pH of the environment and length of the side chain of the phthalate (Wolfe et al. 1980). Details of the degradation pathways of phthalates in landfills have been published in Huang et al. (2013).

Both biodegradation and hydrolysis of phthalates in landfills are influenced by the degradation phase of the landfill, the properties of the phthalate in question, the pH, redox conditions, and organic matter content of the landfill leachate. Due to the low electrophilic nature of phthalate esters, they do not react at neutral pH, but as the pH changes to either acidic or basic, Xu et al. (2008) and Staples et al. (1997) have indicated that hydrolysis is substantially increased. Jonsson et al. (2003a) observed a 99.5% decrease in the concentrations of DMP, DEP, and DBP with initial concentrations of 245 µg/l, 5 µg/l, and 5 µg/l respectively within the first nine months of waste deposition in a landfill whereas a 91.6% increase in DEHP with an initial concentration of 1 µg/l was observed within the same period. These observation on the decrease in DMP and DEP among other phthalates with age, as well as reports by Liu et al. (2010) indicate that hydrophilic PAEs are more easily degraded during the aerobic phase of the landfill than during the anaerobic phase and emphasize the role of the age of the landfill on the types of phthalates contained in its leachate. The specific phthalates identified in landfill leachate could therefore be an indicator of the decomposition phase of the landfill. Fang et al. (2010) and (Huang et al. 2012) highlight the role of the structure of the phthalate in its fate in landfill leachate. They showed that phthalates with short carbon chains such as DMP, DEP and DBP are more easily degraded and mineralised, than the longer carbon chained phthalates like DEHP, which are less susceptible to degradation.

Phenols

Phenols in the landfill are transformed to various compounds through biodegradation and nitrification (Boonyaroj et al. 2017). Nitrification is especially common with the nitrophenols which are first transformed to amino groups under aerobic conditions, and then degraded to CH4 and CO2 under anaerobic conditions (Yazici et al. 2011). According to Sridevi et al. (2012), phenols like the phthalates, can also be degraded under both anaerobic and aerobic conditions, and a positive correlation has been established between oxygen concentration and enhanced degradation of phenols. Microorganisms contributing towards the degradation of phenols include Alcaligenes, Arthrobacter, Pseudomonas, Cyanobacterium, and Bacillus among the bacteria, and Candida, Fusarium, Graphium, Ochromonas, Aspergillus, Phanerochaete, Rhodococus, Rhodotorula, Sphigmonas, and Trichosporon among the fungi and yeasts (Tibbles and Wladyslaw 1989).

Under aerobic conditions, phenolic compounds are oxygenated to catechol through the action of mono oxygenase phenol hydroxylase produced by microorganisms. The catechol is then converted to succinyl Co-A and acetyl Co-A through the ortho pathway of degradation. It can also be converted to pyruvate and acetaldehyde through the meta pathway of degradation depending on the microorganisms present as shown in Fig. 2 (Mahiudddin et al. 2012). Various Pseudomonas species and Alicaligene deutrophus steer degradation through the meta cleavage pathway (Léonard and Lindley 1999), but when Trichosporon cutaneum, Rhodotorula rubura and Acinetobacter calcoacetium are present, cleavage is through the ortho pathway of degradation (Paller et al. 1995). When conditions are anaerobic, phenols are degraded via reduction and benzoic acid formation (Sekman et al. 2011), and these pathways are reportedly much slower and less effective (Yazici et al. 2011; Shibata et al. 2006).

Fig. 2
figure 2

Degradation pathway of Phenols

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For example, under anaerobic conditions, mono and dichlorophenols as well as nitrophenols are degraded less effectively than when conditions are aerobic (Shibata et al. 2006). Among the phenols, the behaviour of BPA in landfills is the most extensively studied and these studies have shown that BPA is non-biodegradable during the aerobic phase of the landfill but as the methanogenic phase begins, fermentation and hydrolysis of wastes produce organic volatile acids, which cause the pH of leachate to decrease and catalyse the degradation of BPA (Zeng et al. 2006; Vieceli et al. 2011, 2014). This emphasizes the role of pH in the degradation of phenols but also indicates that the effect of oxygen on phenol decomposition cannot be generalised.

The role of the structure of the phenol in its fate in landfill leachate was shown with the chlorophenols where the rate of degradation slowed down as the number of chlorine atoms in the phenol increases. Baker and Mayfield (1980) indicated that when chlorine was in the meta-position to the phenolic hydroxyl, the compound was found to be resistant to microbial degradation. Aerobic microorganisms can metabolize mono- and di-chlorophenols, but microbial breakdown of phenolic compounds with a high number of chlorine atoms becomes less effective. In another study by Yazici et al. (2011), the concentration of 2,4-dichlorophenol was found to decrease to below detection levels within 100 days, highlighting the role of the age of the landfill on the degradation of phenolic compounds in landfill leachate. These observations indicate that generalisations on the fate of phenols in landfill leachate may be inaccurate, as several factors may influence their fate in the landfill environment.

Degradation studies on phthalates and phenols have mostly been carried out under carefully monitored environments and preset conditions in the laboratory which may not give a true reflection of what happens in the landfill environment because of the dynamic nature of the landfill conditions, the variety of waste materials that are commonly co-disposed as well as the fact that it is an enclosed environments where each cell may be having a different condition from the next. No two landfills have the same types and quantities of waste received. In addition, the climatic conditions around different landfills vary and so the studies that have been carried out may reflect only landfills with similar conditions. The microbial flora responsible for biodegradation of these compounds in the landfill are also diverse and considering that the landfill is a closed environment where air circulation is limited, the microbial processes are likely to change from aerobic to anaerobic very rapidly and so are the compounds produced from degradation. One can only therefore speculate what happens to these compounds in the landfill as there are several factors interacting with each other to influence the processes that are going on in the landfill environment.

Occurrence of Phthalates and Phenols in Soils Around Landfills

Although steps are usually taken to contain leachate generated in landfills, lack of landfill liners, leaking of liners, and absence of a leachate collection system could result in the spread of leachate into soils in the vicinity of landfills. Even when liners are present and functioning properly, some phenolic compounds are still able to migrate through the liners (Adar and Bilgili 2015). Studies by Makuleke and Ngole-Jeme (2020) and Othman et al. (2019) have shown that leachate may continue to migrate into surrounding soils years after the closure of landfills. Landfills are therefore a major source of pollutants in surrounding soils. Table 4 shows a summary of different phthalates and phenols that have been detected in soils surrounding different landfills around the world.

Table 4 Phthalates and phenols in soils around landfills and their concentration ranges
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The most commonly encountered phthalates in soils around landfills include DEP, DMP,DBP, DnOP, BBP, DPP, DiBP, DnBP, and DEHP (Net et al. 2015). Studies carried out by Gholaminejad et al. (2024) in Iran show that the concentrations of phthalates in soils around landfills followed the order DEHP > DnBP > BzBP > DEP > DnOP > DMP. In another study by Liu et al. (2010) on landfills in China, the landfill did not seem to have a significant effect on the concentrations of phthalates on the top soil but the subsoil had significant amounts of phthalates. Increased amounts of phthalates in subsurface environment could be associated with anaerobic conditions which retard phthalate degradation, or it could be an indication of the migration of phthalates within the subsurface layers of the soils. Phenolic compounds in soils mostly originate from natural sources, as they are present in roots, leaves, and stems of plants, but landfills have been identified as a possible source of phenols as they are contained in leachate from landfills. The concentrations of phenolic compound in soils around landfills vary (Table 4). Nonyl phenols and BPA are among the phenols widely studied in soils.

Studies on the extraction of phenols and phthalates from soils are less reported compared to their extraction from liquid samples but where they have been done, Soxhlet extraction (SE) has been the main extraction technique. According to Singh et al. (2023) Soxhlet extraction is a very efficient technique for the determination of phthalates in soils with recovery rates of between 77 and 121% which is comparable with percentage recoveries of ultrasonic extraction (UE) (75 and 117%) (Tran et al. 2015; Pritchard et al. 2010). Microwave assisted extraction is also used but recoveries are much lower (67–122%) (Liao et al. 2010) than those of SE and UE. In another method by Khosravi and Price (2015) where SPE was used after accelerated solvent extraction (ASE), recovery of spiked samples followed the order DPP (90%) > DMP (89%) > BBP (86%) = DnOP (86%) > DEHP (84%) > DnBP (82%) > DEP (80%) with a limit of quantification of 0.1 ng/kg to 0.87 μg/kg. These recoveries were similar to those of other studies were phthalate recoveries ranged from 17 to 110% (Blair et al. 2009), 86–92% (Reid et al. 2009), 75–120% (Ma et al. 2013) and 86–111% (Khosravi and Price 2015).

Different extractants/solvents have also been widely used in these studies. Khosravi and Price (2015) determined phthalates in soils using methanol and acetonitrile and had percent recoveries of between 36 and 76% with a mean percent recovery of 58%. When acetone/petroleum ether were used as solvent, recoveries of phthalates in soils ranged from 76 to 93% (Chai et al. 2014) which was lower than what was obtained (80–100%) when Mohebbi et al. (2017) used carbon tetrachloride as solvent. Han et al. (2019) reported percentage recovery of phthalates in agricultural soil of 91–107% following extraction with an octanol-based solvent. Soxhlet extraction using solvents such as methanol, acetone, acetonitrile, n-hexane and dichloromethane is the most commonly employed technique for extracting phenols from soils (Santana et al. 2009) and satisfactory recoveries of between 67 and 97% have been obtained with these solvents. A faster extraction method that is also used is UE but it is disadvantaged by the large volumes and cost of organic solvents required. Percentage recovery of phenols using UE ranged between 81 and 99% which is much higher than what is obtained with SE (Li et al. 2004). Combining extraction methods have also proven effective as was observed by Wei and Jen (2003) using microwave assisted extraction and SPME. They obtained recoveries of up to 90%.

Percentage recovery of phthalates in soils are higher when SE and UE and used for extraction compared to microwave procedures. Though the percentage recoveries of the compounds appear to be lower than what is obtained with liquid samples, an acceptable recovery of these compounds in soil is still achievable with modifications. In soils contaminated by landfill leachate which have a cocktail of contaminants, careful attention needs to be given to the solvents used as interference in the extraction of these compounds could be an issue.

Fate of Phthalates and Phenols in Soils

The two main processes determining the fate of phthalates and phenols in soil environments are degradation and sorption (Paszko et al. 2016; Choi and Lee 2017). Other processes such as hydrolysis, transformation into insoluble and recalcitrant humic substances by polymerization and condensation reactions, and dissolution and leaching by percolating water have also been reported especially for phenols (Adeola 2018; Min et al. 2015; Ziolkowska et al. 2020).

Biodegradation and Controlling Factors

Phthalates persist in the soil environment because of the presence of the benzene ring as well as their branching structure and strong hydrophobicity which retards their movement in the soil environment. Their degradation may occur through the biotic and abiotic pathways but the biotic biodegradation is the main degradation route for phthalates in soils because the abiotic pathway is much slower (Singh et al. 2021). Under aerobic conditions, this takes place through sequential hydrolyses of the link between the alkyl chain and the aromatic ring of the diester by esterases produced by microorganisms such as Micrococcus sp. strain YGJ1, Bacillus sp, and Pseudomonas sp, (Çevik et al. 2019; Boll et al. 2020) to produce a monoester and phthalic acid. The phthalic acid is mineralised to either 3, 5, or 4, 5 dihydroxy phthalates, then to procatechuate, followed by pyruvate and oxaloacetate through the ortho pathway, or to acetate, succinate and carbon dioxide through the meta pathway (Cartwright et al. 2000; Singh et al. 2021). When conditions are anaerobic, the phthalic acid is instead converted to benzoic acid, then to catechol, and finally carbon dioxide and water (Staples et al. 1997; Zhao et al. 2016).

With the phenols, biodegradation could also be through the ortho- or meta- cleavage pathways, which are also facilitated by soil microorganisms such as Bacillus sp, Pseudomonas sp, Acinetobacter sp., Achromobacter sp., Fusarium sp., Phanerocheate chrysosporium, Corious versicolor, Ralstonia sp., Streptomyces sp. through the production of specific enzymes including phenol hydroxylase, catechol dioxygenase, and cis-muconate cyclase (Krastanov et al. 2013; Hasan and Jabeen 2015; Nair et al. 2008). Furthermore, phenol hydroxylase breaks down phenolic compounds through dihydroxylation of the benzene ring to produce a catechol derivative, followed by meta oxidation that opens the benzene ring (Hasan and Jabeen 2015). The degradation of the catechol occurs through catechol 2,3-dioxygenase (meta cleavage pathway) or 1,2-dioxygenase (ortho cleavage pathway) as shown in Fig. 2. Oxidization of the catechol produces molecules that can then enter the tricarboxylic acid cycle of the microorganisms. However, not all phenols are amenable to biodegradation. According to Baker and Mayfield (1980), soil microorganisms are able to break down phenol, o-chlorophenol, p-chlorophenol, 2,4-dichlorophenol, 2,6-dichlorophenol, and 2,4,6-trichlorophenol faster under aerobic conditions, than m-chlorophenol, 3,4,-dichlorophenol, 2,4,5-trichlorophenol and pentachlorophenol. Phenolic compounds such as 3,4,5-trichlorophenol and 2,3,4,5-tetrachlorophenol are not susceptible to microbial degradation. Most of this degradation processes occur in soil solution. In landfill environments where there is leachate seepage, the chemistry of soil solution is significantly altered which may affect the congeners arising from the degradation of these compounds in soils. Comparisons between degradation of these compounds in leachate contaminated soils and the same soils which are uncontaminated needs to be studied to understand the effect of leachate on the degradation pathways of these compounds in soils.

Factors Influencing the Degradation of Phenols and Phthalates in Soils

Influence of Compound Structure

The structure of phthalates affects its properties and its degradation in the soil environment. Phthalates with longer chains are more persistent in the soil environment than those with shorter chains, because of increased hydrophobicity with increasing length of the carbon chain and the inability of microorganisms to breakdown long chain phthalates. The branched chain phthalates also degrade faster than the straight chained ones according to Kickham et al. (2012) and this is evidenced by the fact that as the length of the branched chain in phthalate compounds increase, so does their half-life in the soil environment (Zhu et al. 2018). Cousins et al. (2003), have reported half-lives of 1–40, 1–75, 0.4–80 and 25–250 days, for DMP, DEP, DnBP and DEHP respectively in soils. These values are much higher than the half-lives of the respective compounds in water, indicating that the degradation of these compounds in soils may be much slower than in water under aerobic conditions. With regards to phenolic compounds, the form in which they occur in soils plays a more crucial role on their fate than their structure (Schmidt et al. 2011) contrary to what was observed with the phthalates.

Effect of Soil Properties

Soil moisture content, temperature, texture, pH, and organic matter (OM) contents play a significant role in the degradation of compounds in soils because of their influence on microbial activities. Excess or limited water content and low temperatures inhibit microbial activities and could possibly slow down biodegradation of organic compounds in soils. The role of texture in the degradation of compounds in soil could be associated with its influence on soil aeration, water retention, OM content and temperature of the soils, and the level of microbial activities. The effect of texture may vary. Microbial activities are faster when soil texture is silty than when it is clayey or sandy because of the high aeration, and water movement in sandy soils and the contrary in clayey soils. Silty soils present a better environment for microbial activities and could therefore accelerate biodegradation of various compounds in the soils. Soil OM provides the nutrients necessary for microorganisms and so soils rich in OM may have higher microbial flora and activities than vice versa. In landfill leachate contaminated soils however, OM content from decomposing waste is likely to be high because of the high amount of refuse that is rich in OM, especially in municipal landfills. This may enhance biodegradation of compounds. Biodegradation of some phenols however may be limited in soils rich in organic matter because these compounds tend to be sorbed onto humic acids through a non-specific lipophilic action which limits their rate of degradation (Langford and Lester 2002).Soil pH also influence microbial processes and makes available protons that may react with the compounds retarding their degradation.

Sorption of Phenols and Phthalates in Soils and Controlling Factors

The ability of compounds to bind to soil particles vary with their soil sorption coefficients (LogKOC). Yang et al. (2013) indicated that soil sorption coefficient of different phthalates as reflected by their organic carbon/water partition coefficient (LogKOC) values follows the order DMP (1.43) < DEP (1.63) < DMEP (1.95) < DAP (2.25) < DPP (2.34) < BBEP (2.51) < DIBP (2.78) < DBP (3.29) < BBP (3.70) < DCHP (5.0) < DEHP (6.52) < DOP (7.20) < DNP (7.76). This indicates that DMP and DEP would not be easily bound to soil whereas DEHP, DOP and DNP are expected to be immobile because of the high LogKOC values. The mechanism of phthalate sorption in soils has not been widely reported but in a study by Xiang et al. (2019) found that sorption of DBP on soil particles takes place through a three-step process; rapid sorption of the compound on the surface of particles followed by a much slower intraparticle diffusion after the external surface sorption sites have been exhausted, and lastly, an equilibrium stage.

Up to 99% of BPA present in the soil environment is fixed and therefore its mobility from the soil may be limited (Vieceli et al. 2014). A similar situation was observed with nonylphenol which is easily bound to soil organic matter. The adsorption of chlorophenols is correlated with their hydrophobicity and as the hydrophobicity of chlorophenols increases, so does their adsorption whereas adsorption decreases as water solubility increases. Chlorophenols have a LogKOW values of between 2 and 5 and are therefore expected to be soluble. BPA is moderately water-soluble (solubility index is 3.00 × 102 mg/l at 25 °C) and has an LogKOC value of 7519 × 104 (Table 2), hence it is less persistent in the soil environment. However, metabolites of BPA, including 4-hydroxybenzaldehyde, 4-hydroxybenzoic acid 4-hydroxyacetophenone have all been identified in soils (Dodgen et al. 2014). Due to weak acidity, phenols tend to interact with soil minerals through hydrogen bonding and van der Waals forces (Sposito 2016). Sorption of phenols onto soils is however not very significant according to Vidic et al. (1993) because of electrostatic repulsion between the phenol molecules and the negative sites on soil particles, hence they are soluble and very mobile in the soil environment. As with phthalate biodegradation, soil pH, soil texture, temperature, hydrophobicity, and organic matter content among others also affect the sorption of phthalates and phenols in soils.

Effect of Soil pH

An inverse relationship has been reported between sorption of DBP in soils and soil pH (Yang et al. 2013). Increase in pH may decrease sorption of phthalates by soil particles because they get hydrolysed to phthalic acid and protonated phthalic acids in acidic conditions (Gao et al. 2016) (Yang et al. 2013) as was demonstrated with DBP. Soares et al. (2008), Adeola (2018), Ololade et al. (2016) and Choi and Lee (2017) showed that adsorption of chlorophenols, 2,4-Dichlorophenol, nonylphenols, and BPA to soils increases with soil acidity whereas under neutral and basic conditions, it decreases. This behaviour was associated to the protonation that takes place when the soil is acidic, resulting in an increase in the interaction between soil matrix and the phenolic compound. Phenol and phthalate sorption in soil therefore show opposite patterns under acidic pH. Whereas high acidity reduces phenol sorption, it increases that of phthalates. In soils contaminated with leachates from young landfills, phenol sorption is likely to reduce whereas that of phthalates would increase and the reverse may be true for older landfills with alkaline leachates.

Effect of Soil Texture

Contrary to the relationship between phthalate sorption and soil pH, there is a direct relationship between phthalate sorption and soil clay content, as well as between phthalate sorption and organic matter. In a study by Tan et al. (2016), the concentrations of phthalates and their congeners in different soil fractions at a depth of between 0 and 40 cm followed the order 250–2000 μm > (< 2) μm > 2–20 μm > 20–53 μm > 53–250 μm but at lower depths, the pattern was slightly different following the order < 2 μm > 2–20 μm > 20–53 μm > 250–2000 μm. The phthalate distribution pattern in the particles of the surface layers was aligned with the distribution of organic carbon in soil whereas in the lower layers, they were correlated with clay mineral content (Tan et al. 2016; Hwang and Cutright 2003). These observations highlight the role of organic matter in phthalate distribution in soils and also indicate that in the absence of organic matter, clay mineralogy dictates the pattern of phthalate distribution in soils (Xue et al. 2020). Whereas not much has been studied on the role of soil texture and mineralogy on phenol sorption, studies by Ololade et al. (2016) showed that the presence of Mn oxides retards pentachlorophenol (PCP) sorption, while Fe oxides and organic materials both contributed significantly to sorption of PCP which highlights the role of sesquioxides in soils on phenol sorption.

Effect of Soil Organic Matter

The influence of OM on the sorption of phthalates and phenols in soils depends on whether it is dissolved or particulate organic matter present, as well as the specific phthalate or phenol in question. Phthalates have an affinity of DOM and so increase in DOM content in soils would increase their sorption. This pattern is caused by the fact that the carboxyl groups in soil humic acid dissociates at pH 4 and the degree of dissociation increases as the pH of the soil increases resulting in deprotonation or increased negative charges which could result in the humic acids being negatively charged and the phthalates becoming more positively charged due to dissociation in acidic environments which would result in increased adsorption of the phthalates (Wan 2012; Gao et al. 2016). With regards to phenols, not much has been done on its sorption in soils but BPA and nonylphenol mobilities in soil are negatively correlated with organic matter with particle size having a limited impact (Ou et al. 2016). This is contrary to what was observed with the phthalates, where particle size played a major role in its sorption onto soils especially in the absence of soil organic matter (Domene et al. 2009).

Effect of Temperature

Another factor affecting the sorption of phthalates in soils is the temperature (Wu et al. 2015). Sorption of phthalates onto organic matter generally takes place through hydrogen bonding and dipole forces which decrease as temperature increases (Gao et al. 2016). The influence of temperature is also associated with the fact that changes in temperature cause a change in the solubility of the compounds which affects their adsorption. Studies by Gao et al. (2016) showed that the sorption of phthalates like DBP for example decreases with increase in soil temperature and pH indicating that in soils around landfills in the aerobic phase during which the leachate is acidic, the bioavailability of phthalates in the leachate plume may be increased whereas as the landfill moves to the anaerobic phase, their bioavailability could decrease. Whereas temperature may also affect the sorption of phenols in soils, not much is reported in this regard.

Human and Environmental Implications of Phthalate and Phenol in Landfill Environments

The exposure pathways of phthalates and phenols in landfill leachates and surrounding soils to humans are summarized in Fig. 3. Exposure of individuals to phthalates and phenols could have several negative health outcomes including cancer, immunotoxicity, reproductive disorders, and developmental problems (Miettinen et al. 2005; Palmiotto et al. 2014). As EDCs, phthalates affect androgen and oestrogen receptors in males and females respectively (Feng et al. 2020). Testicular dysgenesis syndrome cryptorchidism, hypospadias, undescended testes, reduced anogenital distance, decreased sperm count, increase in sterility and testicular cancer have been highlighted as some of the effects of phthalates such as mono-butyl phthalate (MBP) and mono-(2-ethylhexyl) phthalate (MEHP) in males (Wang et al. 2016; Swan 2008; Benjamin et al. 2017; Lyche 2011). The endocrine effects of phthalates on females include reduced gestational period, delayed puberty in girls, endometriosis low yield of oocytes, and infertility (Feng et al. 2020; Benjamin et al. 2017; Hauser et al. 2016).

Fig. 3
figure 3

Exposure pathways of phenols and Phthalates in landfill to humans

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Long term exposure to phenol, according to Yaqub et al. (2016) may lead to diseases such as cardiac arrhythmias, gastrointestinal damage, liver enlargement, and coma or death. The occupational exposure of phenol has been emphasized. Respiratory diseases such as chronic bronchitis, and cancers like skin cancer, lung cancer, brain cancer and bladder cancer have been reported (Yaqub et al. 2016). Phenolic compounds such as BPA acts as an oestrogen/androgen antagonist, which may reduce sperm count and sperm activity in men (Lee et al. 2003), induce cardiovascular disease and diabetes in both men and women (Huang et al. 2012; Lang et al. 2008), cause dermatitis and contact allergy in both men and women (Tsai 2006), and induce male sexual dysfunction (Li et al. 2010). Other reported endocrine related health complications arising from exposure to BPA include simulation of prolactin release proliferation in a breast cancer cell line, onset of puberty in females and development of male reproductive organs (Markey et al. 2001). The health effects of cresols in humans have also been reported. According to Michałowicz and Duda (2007), exposure to cresols can cause vomiting, hypertension, anaemia, paralysis, tiredness, cramps, and drowsiness in humans. Severe damage to internal organs such as the kidneys and liver by cresol have also been shown (Vanholder et al. 1999). They are also endocrine disruptors, as they are reported to cause infertility in women (De Smet et al. 2003).

Management Implications

Remediation of environments contaminated with phenols and phthalates contained in leachate plumes migrating from landfills requires extensive knowledge of the interactions between the pollutants in the plumes and the soil. Remediation methods used for treatment of environments contaminated with phthalates and phenols include coagulation, flocculation, adsorption, biological treatment, and chemical processes that include ozonation, Fenton process, oncolysis, and photocatalysis among others (Ghosh and Sahu 2022). The success of coagulation/flocculation as a remediation method of phthalates depends on the LogKOW of the phthalate and the presence of organic matter in the leachate as these compounds can be easily sorbed onto OM and together be removed from the water through coagulation/flocculation (Zhang and Wang 2009). Parameters exploited in phthalate absorption include hydrophobicity, structure and polarity of the compound in question, functional groups present, molecular weight, as well as environmental factors such as the prevailing pH, temperature and the properties of the absorbent being used (Tüzün et al. 2005; Rivera-Utrilla et al. 2009; Chen and Chung 2006; Tran et al. 2022). Using biological treatment methods depends on the type of organisms present in the contaminated environments, the oxygen concentration as well as temperature with the solubility of the compound and the length of its alkyl chain playing an important role (Huang et al. 2008; He et al. 2015).

Whereas biodegradation may be used as a remediation strategy, it may be ineffective when long chain phthalates are involved as these organisms are unable to breakdown long chains. Advanced oxidation including ozonation, hydrolysis, Fenton process, sonolysis, and photocatalysis are methods which Zhang et al. (2021) have highlighted could also be used. Details of these processes can be found in Zhang and Wang (2009; Tran et al. (2022), and Ghosh and Sahu (2022). In the soil environment biodegradation, adsorption and partial immobilisation, and chemical oxidation are the main mechanisms used to remediate phthalates (Ordoñez-Frías et al. 2021). These methods however depend on the pH of the soil, the concentration of the compound to be remediated as well as the amount of ultraviolet radiation. Most of the methods could be carried out at small scales or within confined spaces but when large environments are contaminated like the case maybe when leachate plume spreads in surrounding soil environments, environmental factors assume prominence. In the soil environment, the situation may be different as the soil environment is a complex one where biological and chemical processes are taking place simultaneously.

Conclusion

A lot of emphasis has been placed on the occurrence of EDCs in wastewater effluents, but this review has highlighted the occurrence of two EDCs; phthalates and phenols in landfill leachate and soils around landfill environments. The processes in the landfill that affect the occurrence of phthalates and phenols are not fully understood because the landfill is a closed environment with high degree of variability from one landfill to the other and even from one cell to the other within the same landfill. This variability makes generalisations unreliable but inferences from other studies and laboratory simulated studies have shown that octanol water partition coefficient, water solubility index, and vapour pressure of phthalates and phenols influence their concentrations in leachate and their fate in the landfill environments. As the leachate migrates into surrounding soils, properties of the soil influence the sorption and degradation of phthalates and phenols and consequently their fate in the soils. Soil properties of significance in phthalate and phenol behaviour in the soil environment include organic carbon content, pH, redox condition, and textural properties of the soil. More studies need to be carried out on the leaching and degradation of phthalates and phenols around landfills in real life settings because the presence of a cocktail of contaminants in landfill leachates presents landfills and surrounding landfill leachate contaminated soil with chemical conditions that may affect the normal pattern of these compounds in these environments. Generalisations from studies carried out under laboratory settings may result in unreliable conclusions that could compromise management strategies. The complexity of leachate may also affect the recovery efficiencies of various extraction methods and solvents used for analyses of phthalates and phenols in landfill leachates and soils in surrounding environments and so more studies that would identify extraction methods and solvents which are least affected by the complexity of this environment need to be carried out. Accurate presentation of the level of these compounds in these environments and the design of appropriate remediation measures impinge on proper monitoring and availability of reliable information which are influenced by the availability of data and information.