A temporal record of microplastic accumulation in sediment cores of the Great Lakes, North America, reflects macroeconomic and regional influences

Two sediment cores from the Laurentian Great Lakes were studied for MP abundances: (1) LH43, collected from offshore Lake Huron in 2017, and (2) 403A, collected from offshore Lake Ontario in 2017 (Fig. 1). The cores were collected using a box corer by Environment and Climate Change Canada (ECCC) as part of their long-term sampling program. The sediment in each box core was subdivided into 7.62 cm push cores, stored at 4° C on the vessel, and transported back to the lab. Each push core was sectioned at 1 cm intervals and frozen for further analyses. One push core from LH43 was sent to the Institut national de la recherche scientifique for ²¹⁰Pb dating using gamma-ray spectrometry. The ²¹⁰Pb activity profile of the 15 cm sediment core was used to determine the average sedimentation rate and years represented from the surface down, according to the constant rate of supply (CRS) model. A second push core from LH43 was sent to The University of Western Ontario (London, Ontario, Canada) to be assessed for MP content.

A push core from Lake Ontario box core 403A was also sent to The University of Western Ontario for MP investigation. One core sampled in 2006 from the same location as 403A (core 1034) was analyzed by ECCC using the ²¹⁰Po dating technique with alpha spectrometry. The ²¹⁰Po activity profile of core 1034 was used to determine the sedimentation rate and years represented from the surface down, according to the CRS model. To account for the missing ²¹⁰Po data between 2006 (1034) and 2017 (403C), the sedimentation rates were interpolated based on the difference in incremental mass sediment rate between the top two slices of the core. Dating models for cores LH43 and 403A are provided in Table S1.

The laboratory methods used were similar to previous studies that investigated MPs in nearshore and offshore sediment of Lakes Ontario and Erie and in tributary sediment of the Thames River, Ontario (Corcoran et al. 2015, 2020a; Ballent et al. 2016; Dean et al. 2018). Each 1 cm sediment slice was wet sieved using reverse osmosis (RO) water in a 45 µm metal sieve, then transferred into an aluminum pie tray, covered with foil, and dried in an oven at 70 °C. The sediment grain sizes < 45 µm were removed to avoid flocculation during drying in the oven. This step facilitated density separation of MPs and sediment grains but may have resulted in particle count underestimations (Browne et al. 2010; Bergmann et al. 2017). The samples were transferred to glass beakers containing approximately 250 mL of 1.5 g cm⁻³ sodium polytungstate (SPT) solution and were mixed for 2 min using a magnetic stirrer. The supernatant was decanted into a glass separation funnel where it underwent density settling. Once separation was complete, the non-buoyant sediment was removed and the remaining buoyant material, consisting of MPs entrained in organic matter, was filtered through 25 µm quantitative fast flow filter paper and rinsed through a 53 µm sieve to remove any remaining SPT solution. The remaining particles, with a lower limit of 53 µm and an upper limit of 2 mm, were transferred to a glass Petri dish (60 mm × 10 mm) and placed back into the oven at 70 °C for drying.

Anthropogenic (non-natural) particles were identified and classified using a Nikon SMZ1500 stereomicroscope at magnifications of 3.75x to 258x. The particles were classified according to morphology (fibre, fragment, and film), colour, and size. Abundances were reported as number of particles per g of dry weight sediment (N g⁻¹ dw), as opposed to N kg⁻¹ dw which is often used, because the amount of sediment remaining after wet sieving was <1 g (Table 1). Normalizing to number of particles kg⁻¹ dw could therefore result in serious over-extrapolation of MP abundance. The particles were removed from the sample using stainless steel tweezers or dental pick and were placed on double-sided tape that was adhered to a glass slide. Approximately 11% (60 out of 559) of all MP particles were selected for Fourier transform infrared spectroscopy (FTIR) using a random number generator to subtract non-MPs from the pre-FTIR totals. Sixty particles were chosen because of time constraints, as analyzing 559 particles would have taken approximately 55 h. The randomly selected particles were analyzed by FTIR at Surface Science Western, The University of Western Ontario, Canada. Measurements were performed in transmission mode and spectra were collected from 4000 to 600 cm⁻¹ with a resolution of 4 cm⁻¹. Spectra corrections were completed to account for water vapour, carbon dioxide, and the adhesive tape. The spectra were compared with the Bruker spectral library to determine polymer type. The data were calculated and plotted using Microsoft Excel.

In the laboratory, all surfaces and equipment were wiped down with a cotton rag, and glassware was rinsed with RO water and air-dried before use to prevent airborne contamination. Cotton laboratory coats were worn during all experimental phases. A vacuum was used on tile floors to minimize MP particles in dust. Clean-air precautions in the form of a stereomicroscope with metal enclosure, HEPA air purifiers, and glass fibre filters in vents were used during processing and microscopic examination. Three pre-industrial sediment samples from core 403A (25–27, 27–29, and 29–31 cm) were processed and examined to contain zero particle counts. This indicates that the field sampling, as well as processing and examination stages in the laboratory, did not contribute significantly to airborne contamination. Nonetheless, procedural blanks containing pre-examined sediment were placed on the counter during processing and under the microscope enclosure during examination. Airborne contamination was determined from an average of two blanks from each treatment stage. The results were blue (4), pink (2), and black (1) microfibres, which were subtracted from the totals of each sediment sample.

Based on the ²¹⁰Pb results, an average sedimentation rate for LH43 was calculated to be 0.06 g/cm²/year. This indicates that each core slice represents 3.5 years (Fig. 2A). Following FTIR and blank subtraction, the MP abundances in core LH43 ranged from 17.2 to 280.1 g⁻¹ dw (Table 1). Core LH43 exhibits an increase in MP abundance from 14 to 9 cm (Fig. 2A). There is an accumulation peak at 9 cm with a total of 280.1 g⁻¹ dw sediment, a decrease in MP abundance between 9 and 5 cm, and then a general increase from 5 cm to surface. Sediment at 15 cm depth contained an 80% lower MP abundance than that at 1 cm. Core LH43 contained the greatest MP abundances below the surface sediment (>5 cm depth; Fig. 2A). Of 45 particles analyzed by FTIR, the compositions of 14 could not be determined. Of the remaining 31 particles, 76% were determined to be MPs (Fig. 3A). The average fibre abundance in LH43 was 75%, with average fragment and film abundances of 24% and 1%, respectively (Fig. 4A). The polymer types included polypropylene (PP), polyethylene (PE), polyethylene terephthalate (PET), nylon, polyacrylonitrile (PAN), and “paint related” particles (Fig. 3A). Of the analyzed particles, non-plastics included seven cellulose fibres, three polyamide fibres, and one calcium carbonate grain.

The ²¹⁰Po results indicate an average sedimentation rate of 0.027 g/cm²/y for core 1034 (403A). Each core slice thus represents 6.5 years (Table 1). Following FTIR and blank subtraction, the MP abundances in core 403A ranged from 8.2 to 488.4 g⁻¹ dw (Table 1; Fig. 2B). Anthropogenic particles were pervasive in all sediment slices down to 15 cm depth (Fig. 2B). Core 403A displays an increasing MP abundance from 13 to 8 cm. There is an accumulation peak at 8 cm with a total of 285.7 g⁻¹ dw sediment, a decrease in MP abundance between 8 and 6 cm and then a general increase from 6 cm to surface. There are, however, two sharp peaks in abundance within the 5 and 3 cm core slices (Fig. 2B). Sediment at 15 cm depth contained a 96% lower MP abundance than that at 1 cm. The majority of MPs in core 403A were found within the surface sediment (<5 cm depth; Fig. 2B). Thirteen percent of the total number of anthropogenic particles in 403A (15 of 120 particles) were analyzed by FTIR, and 93% were determined to be MPs (Fig. 3B). Fragments were the most common anthropogenic particle type, with an average abundance of 69%, whereas average fibre and film abundances were 31% and 0%, respectively (Fig. 4B). The polymer types found in core 403A included PET, nylon, PAN, PP, polystyrene (PS), PS + nitrocellulose (particle containing both polymers), and polyurethane (PU) (Fig. 3B). Of the particles analyzed, one fragment was of unknown composition.

A chronology of MP deposition for each sediment core is shown in Fig. 2. According to our results, MPs have been accumulating in Lake Huron and Lake Ontario for at least 56 and 72 years, respectively (Table 1; Fig. 2). The results of this study were also plotted against global plastic production (Mt yr⁻¹) (Fig. 5). Although the majority of commercial plastics and resins began to be produced in the 1930s, these products did not begin to be widely produced until the 1950s (Fig. 5; blue curve). The breakdown of plastic items into secondary MPs would therefore have lagged several years behind their production, which is supported by the initial significant increase in MPs into offshore sediment of Lake Huron in 1964 (Fig. 5; orange line) and Lake Ontario in 1974 (green line).

A second lag considered a temporal offset between plastic production and MP accumulation in the environment is evident in Fig. 5. Global plastic production decreased from 51 million tonnes in 1973 to 46 million tonnes in 1975 (Our World in Data 2023). This decrease in plastic production is associated with low global oil supply due to the Arab Oil Embargo (1973–1974). In Canada, plastic product shipments declined during the succeeding 15–20-year period, with a deficit in the trade balance of plastics until 1995 (Government of Canada 2018). A second decrease in plastic production occurred from 1979 (71 million tonnes) to 1980 (70 million tonnes), which can be attributed to the decrease in oil supply during the Iranian Revolution in 1979. In addition, a global economic recession took place from 1991 to 1993. Although this is not evident on the global plastic production curve in Fig. 5, in Canada, the plastic product industry saw a drop in growth trend in the early 1990s (Margeson 2000). This economic recession is considered a result of a rising inflation rate and consequent increase in the Bank of Canada prime rate, in addition to the inception of the Canada–U.S. Free Trade Agreement, which eliminated tariff protection for manufacturing companies. Between 1991 and 1992, average annual growth of employment in the Canadian manufacturing industry fell by 5.7% (Ray et al. 2017), and the gross domestic product (in billions USD) saw a change in growth rate of −2.1% in 1991 (The World Bank 2024).

The effects of the combined periods of decreased plastic production between 1973 and 1993 are represented in Fig. 5 by relatively low abundances of MPs in the cores between 1985 and 2001 (LH43) and 1989 and the mid-2000s (403A). This indicates an approximate temporal offset of 11 years in Lake Huron and 15 years in Lake Ontario between decreased plastic production in the 1970s and later decline of MPs in the environment until the early 2000s. Both cores LH43 and 403A display general increases in MP abundances in the years following 1998–2001, but the profiles are erratic. The abrupt decrease in MP abundance displayed by core 403A in 2010, coupled with the less abrupt decrease in core LH43 in 2012, could possibly represent a third lag effect following the Great Financial Crisis in the years 2007–2009. This global recession, mainly precipitated by a lack of financial supervision by large investment banks and a plunge in U.S. housing markets (Duffie 2019), is clearly seen on the plastic production profile (Fig. 5).

Fluctuations in MP abundances over several decades, such as those identified in the present study, were also found by Turner et al. (2019) in a dated sediment core collected from an urban lake in the U.K. Their results indicate low MP accumulation rates in the 1950s–1960s, then an accumulation peak in the mid to late 1960s, and a subsequent decrease in the 1980s. Regional economic factors have also been shown to have affected MP abundances in sediment cores in China. Mao et al. (2021) identified a rapid increase in MPs between 1978 and 2000 in sediment cores from Wuliangsuhai Lake, China, which they associated with rising industrial levels and significant population growth during that period.

Other investigations of MPs at depth in sediment correlate with worldwide plastic production and population growth. For example, Brandon et al. (2019) identified a substantial increase in MP abundances from 1945 to 2009 in sediment cores from off the coast of Southern California. Similarly, Simon-Sánchez et al. (2022) determined that the mass of MPs increased in sediment cores from the Balearic Sea between 1965 and 2016. Matsuguma et al. (2017) determined that MP pollution has increased over time as recorded in sediment cores from Japan, Thailand, Malaysia, and South Africa.

In a study of freshwater fish from Illinois rivers and Lake Michigan, Hou et al. (2021) quantified MP particles in the digestive tracts of fish caught between 1900 and 2018. In these samples, polyester microfibres were detected but no fibres were identified in fish caught prior to 1950. Averaged fish samples saw a steady increase in MP uptake from 1950 to 2018, which aligns well with general historical trends in cores LH43 and 403A and with global plastics production.

It is possible that the erratic upper parts of the MP profiles, especially for core 403A, could be attributed to the top 6 cm being surface sediment. Although surface sediment in offshore sites is not prone to wave or current reworking, benthic organisms can promote particle mixing through bioturbation. Generally, the displacement of particles through burrowing activities becomes less pronounced at greater depths within a sediment profile (Hülse et al. 2022). Another possibility for the erratic MP abundances in the surface sediment is that smearing could have occurred. Chant and Cornett (1991) experimentally tested the smearing, or mixing of sediment between adjacent core sections, by studying the redistribution of isotopic tracers attached to plastic microspheres during core manipulations. The authors found that 96% of the tracer activity was detected in the upper 2 cm. At depths > 4 cm, tracer activity was <0.1% of the activity in the overlying slices. Adjacent push cores subsampled from individual box core samples were used for the present study, and thus, it is possible that some smearing of the upper slices may have taken place.

Corcoran et al. (2015) studied a core sampled from Lake Ontario in 2013 (403), collected proximal to 403A in the present study. The authors determined that, based on published sediment accumulation rates, MPs had been accumulating in Lake Ontario bottom sediment since ca. 1977. The authors, however, examined sediment ranging from 500 µm to 3 mm in size. In the present study, the MPs examined ranged from 53 µm to 2 mm in size, and this may account for the greater abundances and longer accumulation histories presented herein.

MP abundances in the top 5 cm of core LH43 account for 20% of the total abundance, whereas core 403A contains 61% of its MP abundance in the top 6 cm. These results suggest that over a 15–20-year period (1998–2017), MPs were accumulating at greater rates in Lake Ontario offshore sediment than in Lake Huron. This may be a result of higher inputs from the plastics industry sector, wastewater treatment plants, stormwater drains, and a greater number of people surrounding Lake Ontario (Ballent et al. 2016; Corcoran et al. 2020b; Grbić et al. 2020) (Fig. 1). This regional industry mix effect is supported by a visual inventory of plastic producers and manufacturers that showed 782 plastics-associated companies within 100 km of Lake Ontario, and only 193 within 100 km of Lake Huron (University of Toronto Art Centre 2021).

Anthropogenic particle morphologies differed between cores overall, with LH43 containing 75% fibres and 403A containing 69% fragments (Fig. 3). The average fibre abundance in the top 5 cm of LH43 was lower than the average fibre abundance in the subsurface slices, with 53.2 g⁻¹ dw and 126.5 g⁻¹ dw, respectively, whereas the opposite was seen in core 403A (60.4 g⁻¹ dw in top 5 cm; 24.8 g⁻¹ dw in subsurface slices). The causes of these morphological and depth discrepancies between cores are unknown, but the results suggest that the main sources of MPs and their timing of deposition into each lake differed. This suggests that controls such as population growth, waste water treatment plant development, inception of recycling programs, and plastic industry growth could have been responsible for regional variations in MP abundances over time.

The globally produced polymers, PE, PP, PET, PS, PAN, and nylon are used in a wide variety of plastic products and therefore, their presence in the sediment cores does not indicate specific sources. However, paint-related MPs (epoxy, PU, alkyds, and industrial coatings) compose 40% of all fragments randomly analyzed by FTIR in this study. Some paints are composed of PS and PET, possibly making the percentage even higher in our samples. Paint microchips in lake bottom sediment could be derived from antifouling paints used on boats to deter the encrustation of organisms on hulls. Boucher and Friot (2017) consider marine vessel coatings as one of the main categories of ocean plastic pollution. In addition, a review by Gaylarde et al. (2021) indicates that paint chips are prevalent in global waters, such as the Southern and Atlantic oceans, the Mediterranean and East China seas, and even in Arctic Ocean ice. In the late 1960s, tributyltin (TBT) began to be widely used in antifouling paints, causing serious side effects to non-targeted organisms (Yang and Maguire 2000). Regulations were introduced in Canada in 1989 to limit the use of TBT-containing paints on vessels. Anti-fouling paints containing TBTs were to be removed or encapsulated by 2008 (ECCC 2018), and by 2012, TBT was banned for use and sale in Canada. Nonetheless, the long residence time of TBT in sediment means that paint microchips may remain a threat to aquatic species in the Great Lakes. Bottom sediment samples from the Great Lakes watershed in North America have been shown to contain paint microparticles (Corcoran et al. 2020a; Belontz et al. 2022), and combined with the results of the present study, indicate that paints and other coatings are a significant source of MP pollution.