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Volume 51, Issue 5 p. 780-796
Open Access

The Clean Water Act and biosolids: A 45-year chronological review of biosolids land application research in Colorado

James A. Ippolito

Corresponding Author

James A. Ippolito

Dep. of Soil and Crop Sciences, Colorado State Univ., C127 Plant Sciences Building, Fort Collins, CO, 80523-1170 USA


James A. Ippolito, Dep. of Soil and Crop Sciences, Colorado State Univ., C127 Plant Sciences Building, Fort Collins, CO 80523-1170, USA

Email: [email protected]

Contribution: Conceptualization, Data curation, Formal analysis, Funding acquisition, ​Investigation, Methodology, Project administration, Resources, Supervision, Validation, Visualization, Writing - original draft, Writing - review & editing

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Ken A. Barbarick

Ken A. Barbarick

Dep. of Soil and Crop Sciences, Colorado State Univ., C127 Plant Sciences Building, Fort Collins, CO, 80523-1170 USA

Contribution: Conceptualization, Data curation, Formal analysis, Funding acquisition, ​Investigation, Methodology, Project administration, Resources, Supervision, Validation, Visualization, Writing - original draft, Writing - review & editing

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First published: 26 May 2022
Citations: 1

Assigned to Associate Editor Peter Kopittke.



The 1972 U.S. Clean Water Act set forth the generation of biosolids. In Colorado, biosolids land application research began in 1976 and continues today. Pastureland research suggested that sewage effluent could effectively be land applied to benefit aboveground plant growth and to polish water prior to reaching receiving waters. Forest wildfire–affected ecosystems can also benefit from biosolids applications; application rates of up to 80 Mg ha−1 can lead to greater plant establishment, soil microbial activity, and nutrient turnover and reduced nutrient and heavy metal concentrations in runoff below livestock and USEPA drinking water standards. Long-term (24-yr) observations in oil shale–mined lands showed that biosolids (up to 224 Mg ha−1) can have a positive effect on microbial-mediated nutrient cycling and, in turn, on aboveground plant community structure. Biosolids applications of up to 40 Mg ha−1 in high-elevation shrubland ecosystems, dominated by Mo-containing shale deposits, can aid in reducing imbalances between Mo and Cu in soils and plants; excessive plant Mo, when consumed by ruminants, can lead to molybdenosis. Biosolids and lime applications (both at 224 Mg ha−1) have been shown to improve long-term reclamation success on acid-generating, heavy metal–containing fluvial mine tailings. Thirty years of grazing land research, focused on soil and aboveground plant benefits, illustrate that soil health and plant productivity can be improved to the greatest extent at biosolids application rates close to 10 Mg ha−1. Finally, 40 yr of dryland agroecosystem research (a) have helped identify biosolids N fertilizer equivalency (∼8 kg N Mg−1) and thus dryland winter wheat application rates (e.g., 4.5–6.7 dry Mg ha−1); (b) have identified first-year mineralization rates of 25–32%; (c) dispute the “time bomb” theory by showing that plant metal uptake follows an exponential rise to a maximum; (d) showcase economic return to producers via increased wheat grain protein content; (e) suggest that biosolids-borne proteins and their degradation products are labile C and N sources; (f) have led to long-term tracking of micronutrients and heavy metals in soils and revealed that plants–soil concentrations will not lead to groundwater degradation and that plants are safe for human consumption; and (g) have shown that biosolids provide Zn, helping to overcome soil deficiencies and enhancing Zn biofortification in wheat grain. This latter point is important because ∼2 billion people globally suffer from Zn deficiencies. Forty-five years of research in Colorado have proven that biosolids can enhance environmental quality, improve soil health, and produce healthy food products.

Core Ideas


  • The 1972 Clean Water Act (CWA) governs potential water pollution for public health and the environment protection.
  • The CWA set forth the generation of biosolids.
  • Beneficial use of biosolids has occurred over the past 45 yr in various settings, including pasture and grazing lands, forest fire burned areas, shrubland ecosystems, mined lands, and dryland ecosystems.


  • ammonium bicarbonate-diethylenetriaminepentaacetic acid
  • AM
  • arbuscular mycorrhizae
  • CSU
  • Colorado State University
  • CWA
  • Clean Water Act
  • EC
  • electrical conductivity

    The U.S. Clean Water Act (CWA) of 1972 set forth the USEPA's mission to govern potential water pollution for the protection of public health and the environment (USEPA, 2021b). Among other criteria, the CWA set forth standards that municipalities must meet in terms of cleaning wastewater prior to discharge. Cleaning wastewater leads to removal of solid phases, nutrients, metals, and other chemicals from the water phase and generates biosolids. Subsequently, the beneficial use of biosolids is governed by the CWA 40 Code of Federal Regulations Part 503. This section of the CWA established numerical limits and management practices that protect public and environmental health from potential adverse effects of chemicals present within biosolids during beneficial use such as land application (USEPA, 2021b). The United States currently generates 4.75 million dry Mg of biosolids per year, approximately half of which is land applied (USEPA, 2021a). Although the state of Colorado generates considerably less than the rest of the nation (∼80,000 Mg yr−1 with ∼85% land applied; National Biosolids Data Project, 2018), Colorado has a long history of biosolids land application research.

    Since 1976, Colorado State University (CSU) has been studying biosolids beneficial land application within the state of Colorado, USA. The CSU biosolids land application program, housed within the Department of Soil and Crop Sciences, has spent 45 yr focusing on the benefits of biosolids land application in a number of ecosystems, including (a) pasture and grazing lands, (b) forest wildfire–burned areas, (c) high-elevation shrubland ecosystems, (d) recent and historically mined lands (e.g., oil shale development lands and lands affected by historic precious metal mining), and (e) dryland agroecosystems. This manuscript highlights the systematic use of biosolids within the state within the aforementioned ecosystems, keeping in mind the premise of the 1972 CWA and its rules and regulations with respect to biosolids beneficial reuse and U.S. source water protection.


    2.1 Pasture lands

    The earliest reported biosolids land application research began in Colorado in 1977 and 1978, where Sabey et al. (1981) and Barbarick et al. (1982) studied the effects of sprinkler irrigated sewage effluent or ditch water to a mountain meadow (Fluvaquentic Haploboroll) in Hayden, CO. The site was dominated by a mixture of grasses and forages. At the time, the study premise was to use soil as the final polishing step prior to effluent return to a water source. Sewage effluent or irrigation ditch water were applied at 7.5 cm wk−1 throughout the summer and winter months to identify whether soil could remove problematic substances from the effluent prior to the solution phase returning to sensitive water bodies. Substances of concern included NO3–N, NH4–N, PO4, K, Ca, Mg, Na, Fe, Zn, Cu, Mn, Cl, biochemical oxygen demand and chemical oxygen demand, and fecal and total coliform. Overall, the authors noted little differences between water sources on soil elemental accumulation; only Na, Cl, and PO4 increased slightly with effluent as compared with ditch water. The lack of differences between water sources may have been due to greater aboveground plant yields and subsequently greater total elemental accumulation in plant tissue within effluent-treated meadows. Sabey et al. (1981) further noted that winter effluent application could be effectively used as long as the effluent was applied under ice or snow cover and as long as the soil below was not frozen.

    Column studies by Barbarick et al. (1979, 1980) supported the above research. The authors studied Hayden, CO, wastewater effluent addition on salt flux entering porous cups (measured via electrical conductivity [EC], Na, and Ca contents; Barbarick et al., 1980) or, in a different study, in porous cups, the soil surrounding the cups, or soil within column segments at the end of the study (Barbarick et al., 1979). The soil used was classified as a Fluvaquentic Haploboroll. The overall study goals were to identify sources of error along with a proper soil sampling technique for potential salt accumulation in effluent-treated soils. The authors noted that, although porous cups might be considered a convenient way to monitor soil solution, they tended to either overestimate or underestimate parameters as compared with direct measurements of soil surrounding the cups or via saturated paste extracts.

    2.2 Grazing lands

    Biosolids land application to a cattle-grazed, shortgrass steppe rangeland began in 1991 with a project initiated between CSU and the City of Fort Collins, CO. Biosolids were surface-applied (no incorporation) once at 0, 2.5, 5, 10, 21, and 30 Mg ha−1 to 15 m by 15 m plots on the city-owned Meadow Springs Ranch, Larimer County, CO (40°53′46″ N, 104°52′28″ W; Aridic Argiustoll). The overall project goal was to identify biosolids land application effects on shortgrass steppe plants and soils on the 40,000-ha ranch. Early findings by Barbarick et al. (1992), Harris-Pierce et al. (1993), and Harris-Pierce (1994) showed that increasing biosolids application rates (a) increased plant canopy cover and aboveground warm-season (e.g., dominant plant species included blue grama [Bouteloua gracilis (Willd. ex Kunth) Lag. ex Griffiths] and western wheatgrass [Pascopyrum smithii (Rydb.) Á. Löve]) grass biomass; (b) increased plant macronutrient (e.g., N, P, K) concentrations; (c) did not increase plant metal concentrations above those considered “normal” for plants (Kabata-Pendias, 1984); (d) increased soil NO3–N (1–20 mg kg−1), NH4–N (1–90 mg kg−1), and EC (0.2–1.0 dS m−1) concentrations in the 0-to-8-, 8-to-15-, and 15-to-30-cm depths, with greatest increases occurring in the 0-to-8-cm depth; and (e) increased P, Cu, Zn, Pb, Cd, Mo, and soil organic matter in the 0-to-8-cm depth.

    The Fort Collins–owned Meadow Springs Ranch consists of rolling hills, and thus another concern for the city's biosolids land application program was land application on varying slopes and subsequent effects on runoff water quality. Harris-Pierce et al. (1995) used a one-time application of 0, 22, and 41 Mg ha−1 of biosolids on 8 or 15% slopes (slopes >15% with poor permeability cannot be used based on USEPA [1983, 1995] and Colorado Department of Public Health and Environment [2014] regulations), followed 2 wk later by simulated rainfall at 100 mm h−1 over a 30-min set. The soil was classified as an Ustollic Argiustoll. Runoff was collected and analyzed for sediment content and nutrient/metal concentrations. The only parameter affected by slope was sediment content; greater sediment contents were found on 8% slopes, likely due to greater fine-sized soil particles present as compared with steeper slopes. The remainder of the data was pooled across slopes and showed that, in runoff, (a) Fe and Al concentrations decreased with increasing biosolids application, likely due to complexation with biosolids organic phases; (b) organic N, NH4–N, K, B, P, Cu, Ni, and Mo increased with increasing biosolids application; (c) other elements analyzed (e.g., Pb, Cd, Cr, Zn, Mn, and Ba) were below detection; and, most importantly, (d) all detected elemental concentrations were at least an order of magnitude lower than selected drinking water standards (USEPA, 1992) and recommendations for livestock consumption (Soltanpour & Raley, 1989).

    Core Ideas

    • The 1972 Clean Water Act (CWA) governs potential water pollution for public health and the environment protection.
    • The CWA set forth the generation of biosolids.
    • Beneficial use of biosolids has occurred over the past 45 yr in various settings, including pasture and grazing lands, forest fire burned areas, shrubland ecosystems, mined lands, and dryland ecosystems.

    Understanding that biosolids land application programs eventually re-apply to parcels of land, even in grazed lands, CSU and the City of Fort Collins, CO, partnered in 2002 to re-apply similar biosolids (i.e., both anaerobically digested) at the same application rates to the 1991 plots but to only one-half of each plot. The subsequent research, outlined below, followed a more holistic approach to understanding biosolids land application effects to this ecosystem and subsequently attempted to hone in on a proper land application rate for this and similar locations.

    Sullivan, Stromberger, and Paschke (2006) determined the long- and short-term effects of single or repeated biosolids applications on soil, plant, and microbial community structure. Short-term (i.e., several years after the repeated biosolids applications) soil chemical increases were evident for EC, NO3–N, and NH4–N (0-to-15-cm depth), whereas long-term (i.e., >10 yr after the single biosolids applications) changes were observable only for NO3–N; both short-term and long-term biosolids applications decreased soil pH, affected available P (depending on on-site climatic conditions), and increased total C and N. Furthermore, extractable heavy metal concentrations did not approach USEPA drinking water limits in any plot. The plant community exhibited structural changes in drier as compared with wetter years. In drier years, aboveground plant biomass increased with increasing biosolids application rates, yet plant community richness decreased with increasing biosolids application rates. The soil microbial community exhibited similar structural changes in both wet and dry years, with principle component analysis showing that the 0 Mg biosolids ha−1 separated from the 2.5–10 biosolids ha−1, which separated from the 21–30 biosolids ha−1. The authors performed correlation analyses between all aspects studied, with an outcome suggesting a biosolids land application “sweet spot,” whereby soils, plants, and microbial communities are enhanced to the greatest combined degree may be achieved between the 5 and 10 Mg ha−1 biosolids application rates.

    In a subsequent study at the Meadow Springs Ranch and within the same plots studied by Sullivan, Stromberger, and Paschke (2006), Sullivan, Stromberger, Paschke, and Ippolito (2006) showed that biosolids re-applications temporarily exacerbated soil chemical property differences. However, there were more similarities between short-term and long-term biosolids-amended plots 2 yr after reapplication. Soils that had received 21 or 30 Mg biosolids ha−1 contained greater bacterial biovolumes (i.e., stress indicators) and greater C and N mineralization activity. Microbial communities in soils that received biosolids were able to utilize Biolog substrates to a greater degree as compared with control soil microbial communities. Similar to Sullivan, Stromberger, and Paschke (2006), the authors noted increasing plant biomass but decreasing plant diversity with increasing biosolids application rates both in the short-term and long-term amended plots. Overall results suggested that biosolids land application in this setting, at the rates used, posed little risk of environmental degradation.

    Ippolito et al. (2009, 2014) focused on the fate and transport of Cu and Zn in Meadow Springs Ranch biosolids-amended systems studied by Sullivan, Stromberger, and Paschke (2006) and Sullivan, Stromberger, Paschke, and Ippolito (2006) because Cu and Zn are the greatest biosolids-borne metals applied to the Meadow Springs Ranch. The authors addressed potential downward transport mechanisms and protection against groundwater contamination from both metals. Total Cu and Zn concentrations in soil at various depths (0–8, 8–15, and 15–30 cm) suggested that Zn was immobilized in the soil surface, but findings suggested downward Cu transport. The authors performed subsequent studies using a sequential fractionation procedure by Sloan et al. (1997), chemical equilibria modeling, and scanning electron microscopy with energy dispersive spectroscopy. Findings supported the fact that the single or repeated biosolids applications did not affect downward Zn movement. However, downward Cu transport appeared to be associated with organically complexed phases. Given time, these organically complexed Cu phases were transformed, likely via microbial activity, to create insoluble Cu–oxyhydroxide associations. Overall, soluble Zn and Cu in this biosolids-amended system were one to three orders of magnitude lower than USEPA drinking water standards, and thus biosolids land application in this system would not lead to environmental degradation with respect to Cu or Zn.

    The Meadow Springs Ranch site studied by Sullivan, Stromberger, and Paschke (2006); Sullivan, Stromberger, Paschke, and Ippolito (2006); and Ippolito et al. (2009, 2014) was revisited in 2018 to perform a soil health assessment of the single or repeated biosolids plots. Buchanan and Ippolito (2021) used the Soil Management Assessment Framework to determine soil physical, chemical, nutrient, and biological characteristics and an overall soil health index. The authors found no change in physical (e.g., soil bulk density and aggregate stability) soil health because this was maximized on-site regardless of biosolids application. However, soil chemical (pH, EC), nutrient (available P and K), biological (soil organic C, microbial biomass C, potentially mineralizable N, β-glucosidase activity), and overall soil health were maximized at 0–21, 0–10, 30, and 0–21 Mg ha−1, respectively. Using a conservative yet holistic approach, these findings suggested that applying biosolids at ∼10 Mg ha−1 every 12 yr to this site would be a reasonable target application rate. This finding also supports the “sweet spot” land application rate approach suggested by Sullivan, Stromberger, and Paschke (2006). Finally, in late 2021, the City of Fort Collins, CO, used the above results to establish a new set of research plots that bracket this “target” application rate. Five replicates of biosolids were applied at 0, 5, 10, and 15 Mg ha−1 to 25-m-wide by 900-m-long plots, with the future goal (if funding becomes available) of identifying a true biosolids application rate for beneficial reuse and environmental protection in this shortgrass steppe ecosystem.

    2.3 Forest wildfire–burned lands

    Historic settlement in the western United States led to fire suppression in the early 20th century. In turn, fire suppression has led to catastrophic fires throughout the western United States over the past several decades. One such fire, the 1996 Buffalo Creek fire, which occurred ∼50 km south-southwest of Denver, CO, was a high-intensity, fast-moving, stand-replacing crown fire that burned ∼5,000 ha (Meyer et al., 2004). This fire removed surface vegetation and litter, created an impervious hydrophobic layer, and exacerbated sediment erosion. Unfortunately, this watershed provided drinking water to the city of Denver, CO, with a population of ∼ 2,000,000 at the time of the fire. In an effort to reduce erosional losses, CSU worked closely with the U.S. Forest Service, the USGS, and the Denver Metro Wastewater District to reduce sediment and nutrient loads moving off-site and into Denver's drinking water sources. Various rates of composted biosolids were applied ∼1 yr after the wildfire to soils classified as Typic Ustorthents, Typic Ustochrepts, and Typic Haploborols, and two studies were performed as outlined below.

    Meyer et al. (2001) used a simulated rainfall experiment, studying the effects of sediment and nutrient loss due to increasing composted biosolids rates (0, 40, and 80 Mg ha−1) applied to paired plots on 10–16% slopes. Two years after composted biosolids applications, artificial rainfall was applied at ∼100 mm h−1 for 30 min, and water samples were collected in downslope troughs. The authors noted that sediment concentrations were always greatest in the control and least in the 80 Mg ha−1 application rate. Mean runoff concentrations of Ca, Mg, Al, Mn, Sr, Ba, and Si were always greater in the controls as compared with composted biosolids-treated plots, attributed to element-associated mineral (i.e., sediment) transport. However, runoff Na, K, P, and Cu concentrations were always greater in the plots treated with 80 Mg ha−1, attributed to greater concentrations of these elements present in the composted biosolids (Meyer et al., 2001). The authors also noted that Fe, Zn, Ni, Mo, Cd, Cr, As, Se, and Hg runoff water concentrations were similar across all treatments. Lead (Pb) runoff water concentrations were always lowest in the 80 Mg ha−1 rate, likely due to Pb sorption via the biosolids matrix (Hooda & Alloway, 1993). Based on these results, Meyer et al. (2001) concluded that the mean runoff elemental concentrations were less than those recommended for livestock drinking water (Soltanpour & Raley, 1989) and less than the USEPA drinking water standards (USEPA, 1992), except for Pb at the 0 and 40 Mg ha−1 application rates.

    In a subsequent study, Meyer et al. (2004) studied the effects of increasing composted biosolids application rates (0, 5, 10, 20, 40, and 80 Mg ha−1) on plants and soils over the 4 yr after application at the Buffalo Creek burn site. After composted biosolids application, all plots were hand seeded with a mixture of native grasses (see Meyer et al. [2004] for specific species) at 27 kg ha−1, followed by dragging a weighted chain link fence across all plot surfaces to cover the seed. Over the 4-yr study, total plant biomass and bare ground generally increased and decreased, respectively, with increasing composted biosolids application rates. Greater composted biosolids application rates also increased plant N, P, and Zn. Shortly after composted biosolids application (2 mo), total soil C and N increased with depth (down to 30 cm), but this effect was inconsistent over the 4-yr study; composted biosolids mineralization via enhanced microorganism activity must have occurred. Overall, the findings of Meyer et al. (2001, 2004) suggested that increasing composted biosolids application rates can aid in forest wildfire–burned area rehabilitation by increasing plant establishment, improving belowground microorganism activity that leads to improved nutrient cycling and aboveground plant uptake, and reducing runoff nutrient losses to below those considered hazardous to livestock and humans.

    2.4 High-elevation shrubland ecosystems

    High-elevation ecosystems of Colorado are those found at relatively high elevation, receive relatively low precipitation, and are dominated by sagebrush (Artemesia spp.). Past evidence in these types of ecosystems suggests that when water availability is adequate, soil nutrient availability may limit aboveground plant growth (Breman & deWit, 1983). Thus, biosolids may be a viable land application option to improve aboveground plant growth and belowground soil constituents in these ecosystems.

    Pierce et al. (1998) surface applied biosolids (0, 5, 10, 15, 20, 25, 30, 35, and 40 Mg ha−1) to soils (Aridic Argiborolls and Borollic Haplargids) in a sagebrush-dominated community near Wolcott, CO (2,225 m asl; mean annual precipitation, 356 mm). At this site, oil shales below the soil surface tended to contain elevated Mo concentrations, leading to imbalances in Cu and Mo in aboveground plants that may lead to molybdenosis in browsing animals. The project goal was to identify biosolids application effects on canopy cover, aboveground biomass, and overall nutrient status (i.e., in additional to other elements and Cu and Mo ratios) in aboveground plants over several years after land application. In general, during years that received greater than mean annual precipitation, increasing biosolids application rates led to greater aboveground biomass and forage tissue N concentrations as well as improved plant tissue Cu/Mo ratios to the extent that forages were safer for ruminant consumption. Similar to the conclusions of Sullivan, Stromberger, and Paschke (2006), Pierce et al. (1998) identified a biosolids application rate “sweet spot” of between 10 and 25 Mg ha−1, whereby plant biomass was maximized and plant nutritional improvements occurred to the extent that they would be safe for animal (e.g., wildlife, livestock) consumption.

    In a subsequent study, Barbarick et al. (2004) researched the effects of the 0 as compared with the 40 Mg ha−1 biosolids application rates (i.e., from Pierce et al., 1998) on soil microbial activity in the sagebrush-dominated community near Wolcott, CO. Soils were obtained 6 yr after biosolids application and examined for microbial respiration, potential N mineralization, arbuscular mycorrhizae (AM) fungi root colonization, and substrate-induced microbial respiration. Barbarick et al. (2004) found that, as compared with the control, the 40 Mg ha−1 biosolids application rate increased (a) microbial respiration 2.3-fold, (b) potential N mineralization 5.4-fold, (c) AM fungal root associations by 33%, and (d) substrate induced microbial respiration by 13%. These relatively short-term effects (i.e., 6 yr after biosolids application) after biosolids land application suggested that labile C and N were still present and being utilized by the soil microbial population. The presence of C and N likely enhanced AM fungal associations with plant roots, along with greater quantities of active microbial biomass. These results are similar to those of Sullivan, Stromberger, and Paschke (2006) and Sullivan, Stromberger, Paschke, and Ippolito (2006), suggesting that biosolids can enhance both the short-term (i.e., several years) and long-term (i.e., decadal) enhancement of soil microbial activity enhancement in high-elevation sagebrush ecosystems.

    2.5 Reclamation of mined lands

    In the United States, federal law prohibits the use of biosolids at rates that exceed agronomic requirements of crops, except in land reclamation activity cases and with written consent of a regulating authority (McFarland et al., 2010). The following section focuses on the use of biosolids in reclamation activities within mined lands throughout Colorado and at rates much greater than those typically applied to agronomic crops.

    2.5.1 Oil shale development lands

    In the Piceance Basin of northwestern Colorado (2,040 m asl; mean annual precipitation, 325 mm), high-elevation sagebrush ecosystems overlay oil shale deposits. This area contains a reported 1,200 billion barrels of shale oil, enough to supply the U.S. energy needs for centuries (USGS, 1987). Prior to development, local Ute Native Americans termed rocks in the area as “the rock that burns”; mining claims and oil shale mining began in the early 1900s (USGS, 1987). As mining operations became larger, the USGS (1987) suggested that the amount of waste rock generated per day could be in the millions of cubic meters, eventually leading to several oil shale mine reclamation projects involving the use of biosolids.

    In 1976, one of the first projects using biosolids as a substrate for oil shale land reclamation began in Colorado. Sabey et al. (1980) reported the short-term (i.e., 2 yr after application) effects of biosolids plus wood chips (56 + 0, 112 + 22.4, and 224 + 44.8 Mg ha−1 of each material, respectively), compared with comparable inorganic fertilizer applications (control), on aboveground and belowground ecosystem alterations in disturbed oil shale lands. Out of all biosolids treatments, the lowest biosolids application rate produced the greatest aboveground plant biomass and cover. In general, all biosolids application rates appeared to cause no detrimental short-term effects in aboveground plant growth. Biosolids benefits may have been due to microbial mineralization of both organic N and P compounds present in biosolids or to biosolids altering soil physical properties that affected soil moisture retention, temperature regulation, and aeration. Sabey et al. (1980) also found that soil EC was greatest in the plots treated with 224 Mg ha−1 biosolids (6.6 dS m−1) and extractable P was greatest with the plots treated with 112 and 224 Mg ha−1 biosolids as compared with the control. Other soil constituents measured (pH, NH4–N, NO3–N, total N) were not statistically different between any treatments. In particular, the N findings perplexed the authors.

    In a subsequent study, Voos (1984) attempted to identify the nuances of N transformations within the system studied by Sabey et al. (1980). The author performed a 4-mo, destructive sampling incubation study using biosolids from the same source as Sabey et al. (1980) (i.e., from Steamboat Springs, CO) at rates of 0, 40, 80, and 120 Mg ha−1 applied to topsoil (0-to-15-cm depth; Lithic Croboroll and Pachic Paleboroll) and coal mine spoil. Voos (1984) observed increases in both soil NO3–N and NH4–N concentrations over time, with topsoil amended with biosolids always containing greater NO3–N and NH4–N. Greater inorganic N concentrations in topsoil-amended soils were likely due to enhanced N mineralization in topsoil, although the author noted that ammonification increased with increasing biosolids application rates over time. However, nitrification did not statistically change over time; this was attributed to lower quantities of nitrifying bacteria present or to a detrimental effect on microorganisms due to potential increases in NH3 concentrations. Overall results suggested that NO3–N leaching would likely not be an issue in these coal mine–affected areas; yet these in-laboratory observations still did not support the short-term soil N observations of Sabey et al. (1980).

    Because biosolids land application to high-elevation, sagebrush-dominated ecosystems requires a more long-term ecological context, Paschke et al. (2005) studied biosolids land application effects on soils (Borollic Camborthids) in a disturbed sagebrush steppe ecosystem in Piceance Basin ∼65 km northwest of Rifle, CO. The authors focused on soil fertility and plant community composition 24 yr after biosolids application (0, 56, 112, and 224 Mg ha−1) to two sagebrush-steppe soil materials disturbed by oil shale development (replaced subsoil: initial low soil fertility; topsoil over shale: initial high fertility). Study results suggested that long-term microbially mediated nutrient cycling, due to biosolids-borne organic substrate addition, appears to be favored in biosolids-amended plots, similar to results observed by Pierce et al. (1998). However, Paschke et al. (2005) found that the long-term effects of biosolids application on plants depended on the original substrate that was amended with biosolids. Biosolids-amended subsoils (i.e., stockpiled subsoil amended with biosolids and then placed over existing, on-site soil substrate) showed a reduction in plant diversity; perennial grasses dominated the ecosystem. Biosolids-amended topsoils (i.e., stockpiled topsoil amended with biosolids and then placed over retorted oil shale) showed a decrease in perennial grasses and a subsequent increase in sagebrush dominance. The authors made a significant note on the plant community findings: “If the goal of restoration is to restore a sagebrush steppe community, then the use of biosolids can be viewed as having a negative effect on nutrient-poor soils in the long-term. However, biosolids use on nutrient-rich soil can be viewed as having a positive effect in the long-term” (Paschke et al., 2005) in terms of restoring this sagebrush plant community.

    2.5.2 Heavy metal–contaminated lands from historic mining operations

    Summitville Mine

    The Summitville Mine in southwestern Colorado originally contained an abundance of sulfide-bearing minerals and thus had been the target of underground gold mining from the 1870s through the 1950s, followed by open-pit gold mining from 1984 to 1992 (Rieder et al., 2013). In 1994, this site was declared a Superfund Site by the USEPA, which provides funds and grants the USEPA authority to clean up contaminated sites (USEPA, 2021c). The site posed a number of challenges for reclamation, including high elevation (3,500 m asl, with a mean annual precipitation of 1,400 mm received mostly as snow) and relatively short growing seasons, high acid production potential, low organic matter content, and low soil microbial activity within on-site waste materials (Sydnor & Redente, 2002).

    In 1995, Rieder et al. (2013) began a greenhouse trial to identify promising treatments to be used in a subsequent field study. Waste rock (pH 2.8) and two topsoils (undisturbed topsoil from an area near the mine and stockpiled during mining operations) from the Summitville Mine were returned to CSU. Waste rock amendments included the two topsoils separately or applied in a 50:50 blend; organic material additions that included manure, mushroom compost, biosolids (all applied at 90 Mg ha−1); wood waste (45 Mg ha−1); and various acid-neutralizing materials (∼31 Mg lime 1,000 Mg−1 waste rock or ∼9 Mg lime 1,000 Mg−1 stockpiled topsoil). A 4-mo greenhouse study was established to identify plant characteristics such as aboveground biomass. Results showed that (a) stockpiled and non-stockpiled limed soil produced similar aboveground biomass and (b) mushroom compost and manure produced similar aboveground biomass, with biosolids producing slightly less biomass.

    The above greenhouse trial led to the field trial by Sydnor and Redente (2002). The researchers began an on-site study, incorporating combinations of limestone (102 Mg ha−1) and organic amendments (mushroom compost at 0, 90, and 135 Mg ha−1; biosolids applied at 0, 90, and 135 Mg ha−1) in waste rock, overlain with limed (8.3 Mg lime 1,000 Mg−1 acidic topsoil; pH 2.8) or unlimed topsoil. The project objective was to identify material combinations that would promote aboveground biomass and lessen trace element uptake by perennial plants (i.e., a seed mixture of grasses and forbs). Short-term findings (i.e., 2–4 yr after material applications) showed that (a) both mushroom compost and biosolids treatments supported greater aboveground plant growth compared with treatments that did not receive an organic amendment (the mushroom compost was more effective than biosolids, potentially due to greater nutrient content); (b) plant Cu, Cd, Zn, and Pb concentrations were below those considered phytotoxic (although some plant samples contained elevated Mn concentrations, no visual Mn toxicities were observed); and (c) total soil Cu, Zn, Cd, and Pb were elevated in some plots, yet no plant toxicity symptoms were noted. The overall outcomes of this short-term field study pointed toward expectations that future plant species composition and aboveground biomass would change as plants adapt and modify the environment and as soil development proceeds (Sydnor & Redente, 2002). Keeping this notion in mind, along with the above research outcomes, a site-wide restoration project began in 1999. Mushroom compost, agricultural-grade lime, and stockpiled topsoil were selected for this new project based on the positive effect they had on total biomass production (Rieder et al., 2013). Biosolids was not chosen, proving that, although useful, it may not always be the best solution.

    Leadville Mining District

    The Leadville Mining District in central Colorado began in 1860 and has included four mining activity stages: (a) gold placer, (b) silver lodge, (c) gold lode, and (d) base metal extraction (Colorado Preservation, 2022). Historic mining operations have contributed to metals contamination in soil, sediment, groundwater, and surface waters, leading to (among other issues) acid-generating alluvial mine tailings deposits along stretches of the Arkansas River headwaters (Figure 1, inset). Like the Summitville Mine, this area was listed as a Superfund Site (USEPA, 2022) and poses similar reclamation challenges (3,100 m asl; mean annual precipitation, 310 mm, mostly as snow; short growing seasons; mean days with freezing temperatures, 278 d). This area has received the most attention with respect to biosolids land application use for mine land reclamation within Colorado.

    Details are in the caption following the image
    Leadville, CO, alluvial mine tailings reclaimed with biosolids and lime both at 224 Mg ha−1 in 1995. Background picture from 2019 (courtesy of Jim Ippolito). Inset picture of the site prior to reclamation showing surface Zn salt precipitates (as great as 9%) (National Academy of Sciences, 2003)

    Brown et al. (2005) began a study in 1998 focused on biosolids and limestone application (both applied at 224 Mg ha−1; incorporated to 20 cm) to a highly acidic (pH 3.4), metal-contaminated alluvial mine tailing near Leadville, CO. This mine tailing was initially devoid of vegetation. The goal was to quantify short-term ecosystem functionality after biosolids and limestone application. The authors noted significant decreases in extractable Cd, Pb, and Zn in biosolids + limestone as compared with untreated alluvial mine tailings (e.g., water-extractable Cd concentrations decreased from ∼10 to ∼0.1 mg kg−1). Brown et al. (2005) also observed increased microbial activity in biosolids + lime–amended tailing via increases in CO2 evolution and NO3 generation and reductions in N2O emissions. The authors performed survivability studies with ryegrass (Lolium perenne L.) and earthworms (Eisenia foetida), noting that both species died in unamended tailings. Amendments led to lower Cd, Pb, and Zn accumulation in earthworms and to either trends or significant decreases in Cd, Pb, and Zn within ryegrass. Regardless, biosolids + lime did not reduce ryegrass Pb and Zn concentrations to those below potentially considered phytotoxic, although phytotoxicity was not observed and plant growth was not suppressed. Brown et al. (2005) also performed work with on-site animals, noting that (a) amended tailings can potentially reduce Cd and Pb accumulation in small mammals, and (b) potential resuspension of amended tailings into water bodies did not affect fathead minnow (Pimephales promelas) survivability (>90%), whereas resuspension of unamended tailings caused 100% mortality. Brown et al. (2005) concluded that short-term ecosystem functionality had been restored, but the system had not yet reached equilibrium; obviously, long-term data are required to ascertain prolonged positive effects of biosolids land application.

    To further understand potential reclamation success within the Leadville Mining District, Svendson et al. (2007) performed column studies where alluvial mine tailings were amended with biosolids (224 Mg ha−1) and five different liming materials (coarse-textured lime, agricultural lime, fine-textured lime, sugar beet lime, and kiln dust lime; all applied at 224 Mg ha−1 CaCO3 equivalent) and overlain on top of unamended tailings; a control and the fine lime alone were also used as treatments. All treatments increased alluvial mine tailings pH from about 4.0 to near neutral. Treatments containing biosolids also aided in raising tailings pH below the zone of incorporation, and as a result (a) decreases in extractable Zn and Cd were noted in both the treated zone and in the untreated submaterial, and (b) leachate Cd and Zn concentrations were reduced by up to approximately one to two orders of magnitude (Cd: from ∼5 [control] to between 0.1 and 1 mg kg−1; Zn: from ∼500 [control] to between 5 and 100 mg kg−1). Svendson et al. (2007) suggested that the most successful amendments were biosolids co-applied with either sugar beet lime or kiln dust lime, which are waste products from other industries.

    In 2000, Brown et al. (2007) established field plots in alluvial mine tailings in Leadville, CO, in an area that had been devoid of vegetation for over 70 yr to ascertain the effect of altering amendment C/N ratios on the aboveground plant community status. Based on the results of Svendson et al. (2007), kiln dust lime (224 Mg ha−1) was added to all plots, and then biosolids and woody debris were co-applied at ratios that increased the target C/N ratio from 8:1 to 50:1 (for more detailed information, see Table 1 in Brown et al. [2007]). A native seed mixture was then hand applied and raked into the soil. Total soil C, soil pH, metal toxicity, plant cover, and plant species numbers were determined over the subsequent 5 yr. Brown et al. (2007) noted that (a) total soil C decreased over time, likely related to increased microbial activity over time; (b) all amendments reduced soil metal toxicity due to increasing pH, leading to (c) increased plant cover over time with all treatments; (d) increasing C/N ratios tended to increase plant species richness; and (e) amendments mixed at C/N ratios >20:1 appear to increase plant species diversity to the greatest extent over time.

    Understanding the concerns of Brown et al. (2005) that long-term data are required to ascertain prolonged positive effects of biosolids land application, the initial site established in 1998 by Brown et al. (2005), using 224 Mg ha−1 of both biosolids and limestone amended into highly acidic, metal-contaminated alluvial mine tailings, was revisited between 2006 and 2007 by Freeman (2008) and Freeman et al. (2008). The authors used a number of techniques to quantify long-term reclamation success. Sequential trace metal extractions showed that, out of all metals, Cd was of the greatest bioavailable concern because it was dominantly found within relatively mobile soil phases. Within on-site areas that contained relatively high metal availability, microbial activity was relatively low (quantified via a dehydrogenase assay). These areas of high metal availability also had indicators of microbial community stress (e.g., greater presence of Gram-negative bacteria). Before reclamation, this site supported no vegetation; however, ∼10 yr after amendment applications, Freeman (2008) noted that plant cover equaled ∼73%, litter comprised ∼23%, and bare ground comprised only ∼5%, indicating reclamation success; Ippolito noticed similar reclamation success in 2019 (Figure 1). Plant heavy metal uptake tended to be greatest in areas indicative of microbial community stress. Freeman (2008) correlated Mehlich-3 soil extractions with metals associated with the soluble/exchangeable and weakly bound phases and related these finding to plant metal uptake and overall ecosystem health. Overall, on-site metals had been transformed to more immobile phases, yet issues were still present in areas of relatively high metal bioavailability. Freeman (2008) suggested that a simple Mehlich-3 extraction could be used to identify areas in need of future reclamation.

    Other biosolids research conducted in the Leadville mining district included:
    • A 2-mo greenhouse study focused on riparian shrubs: thinleaf alder (Alnus incana, spp. Tenuifolia), water birch (Betula occidentalis Hook.), red osier dogwood (Cornus sericea L.), shrubby cinquefoil [Dasiphora fruitacosa (L.) Rydb. ssp. floribunda (Pursh) Kartesz.], and Geyer willow (Salix geyeriana Andersson). Survivability in three alluvial mine tailings amended with composted biosolids (224 Mg ha−1) and variable lime rates (3.5–17.2 Mg ha−1) (Davis et al., 2008). All species survived the experiment, with total biomass increasing between 525 and 831%. Davis et al. (2008) noted that all five species had the potential for reclamation use because they had the ability to exclude Pb and Zn below values consider toxic to wild and domestic animals. Furthermore, establishing these shrubs on-site may significantly reduce the extent of waterways reported to be affected by inactive mining and subsequent metal-containing sediment transport, as described under section 305(b) of the CWA (Davis et al., 2008).
    • A follow-up greenhouse study to Davis et al. (2008) by Meiman et al. (2012). Additional plant species were used, with relatively similar outcomes to the Davis et al. (2008) study.
    • A field study focused on survivability of either pre-rooted or staked (i.e., cuttings from) Geyer willow and mountain willow in lime (100 Mg ha−1) and composted biosolids–amended (200 Mg ha−1) tailings (Bourret et al., 2009). Although lime addition reduced tailings metal bioavailability, regardless of planting method, both willow species concentrated Cd, Mn, Pb, and Zn in leaf tissue above levels considered toxic to plants. After 4 yr, pre-rooted mountain willow had greater survivability, suggesting that this species may be used for reclamation. However, care should be taken to not overplant this species because it did accumulate Cd and Zn to concentrations great enough to potentially cause animal and wildlife toxicity if ingested in large enough quantities (Bourret et al., 2009).
    • A companion greenhouse study to the field study of Bourret et al. (2009), whereby Geyer and mountain willows were grown in topsoil versus lime (100 Mg ha−1) plus biosolids–amended (224 Mg ha−1) tailings (Boyter et al., 2009). Geyer willow grew to a greater extent in topsoil, but soil type did not affect mountain willow growth. Geyer willow contained greater Mn, Pb, and Cu concentrations, whereas mountain willow contained greater Cd concentrations. Observations and conclusions were similar to that of Bourret et al. (2009).
    • A field study that followed up on the greenhouse study of Svendson et al. (2007), whereby Brown et al. (2009) amended tailings with 224 Mg biosolids ha−1 and 224 Mg ha−1 of either sugar beet lime, agricultural lime, or kiln dust lime. The goal was to identify which combination would best restore vegetative cover. Unfortunately, limited precipitation increased volunteer species, and the authors suggested that supplemental irrigation may be needed to improve overall reclamation success.

    2.6 Dryland agroecosystems

    The largest body of biosolids research in the state of Colorado encompasses land application to dryland agroecosystems in eastern Colorado. Dr. Ken Barbarick initiated biosolids land application trials near Bennett, CO, in 1982, with support from the then Littleton/Englewood wastewater treatment facility (now entitled South Platte Renew). Yearly funding has been supplied by this facility every year since 1982, making this research program 40 yr old as of 2022.

    2.6.1 East and West Bennett, CO, biosolids land application research program

    Beginning in the 1982–1983 and 1983–1984 cropping years, liquid biosolids (35–42% solids; Figure 2) were applied once to a total of four (two per year) dryland wheat–fallow plots at rates of 0, 6.7, 13.4, 27, and 40 dry Mg ha−1. Comparable N application rates were also used (urea; 0, 34, 67, 101, and 134 kg N ha−1). All materials were fully incorporated into the top 20 cm of soil (soils at all sites were Aridic Paleustolls). Utschig et al. (1986) were the first to report the short-term (i.e., 1 yr after application) findings from this research. Increasing biosolids land application rates increased winter wheat yield and protein content; biosolids also improved yield and protein content over inorganic N fertilizer applications. Wheat grain protein content was >12% in all biosolids-treated plots; baking qualities of wheat increase at or above 12%, and millers often pay more for such wheat (Utschig et al., 1986). Greater wheat grain protein content favored greater economic return to producers as compared with N fertilizer applications. Biosolids-borne heavy metal (Zn, Pb, Cd, Cr, Cu) uptake by plants was relatively low, likely due to sorption by soil, biosolids-borne organic matter complexation, or soil mineral precipitation (Utschig et al., 1986). It is interesting to note that Zn deficiencies can occur in calcareous soils such as those found in eastern Colorado (Westfall & Bauder, 2011); biosolids contain Zn, which could act as a Zn fertilizer. Soil and wheat grain Zn issues, in conjunction with biosolids land application, will be revisited often in this section.

    Details are in the caption following the image
    Liquid biosolids, from the Littleton/Englewood wastewater treatment facility, applied to research plots in 1982 near Bennett, CO (courtesy of Utschig et al. [1986])
    The plots described by Utschig et al. (1986) continued to receive biosolids every other year from 1982 through 2002 simply due to the cropping rotation being wheat fallow. However, biosolids were air-dried in sand drying beds prior to on-site delivery where the material was hand applied, raked to uniformity, and then roto-tilled into the top 20 cm of soil (Figure 3). Subsequent research on these plots focused on repeated biosolids application effects on:
    • Heavy metal and soil NO3–N accumulation. Lerch et al. (1990a) showed that increasing biosolids rates resulted in increases in surface soil (0–20 cm) extractable Zn, Cu, Cd, Ni, and Pb concentrations as compared with the control. The most limiting metal was Cu, which, when applying biosolids at an agronomic rate, would reach USEPA maximum cumulative loading in 144–456 yr based on biosolids applied every other year to a wheat–fallow rotation. Lerch et al. (1990a) also noted that excessive biosolids applications (e.g., 27 Mg ha−1) increased soil NO3–N concentrations (20–28 mg kg−1) to a depth of 150 cm; these concentrations are greater than winter wheat crop requirements. Lerch et al. (1990a) were the first to recommend 6.7 dry Mg biosolids ha−1 as an agronomic rate, where soil NO3–N concentrations were comparable to controls. This application rate would also help protect against groundwater contamination if groundwater was relatively close to the soil surface.
    • Winter wheat production and producer income. Lerch et al. (1990b) observed little differences in wheat yield between increasing biosolids or inorganic N fertilizer applications. However, identical to the findings of Utschig et al. (1986), the authors noted that biosolids resulted in greater economic return to the producer as compared with inorganic N applications, primarily driven by increases in grain protein content and protein premiums being paid by millers. Lerch et al. (1990b) noted that the suggested agronomic rate of 6.7 dry Mg biosolids ha−1 would return an average of $111 ha−1 as compared with comparable inorganic N fertilizer applications due to increased grain protein content.
    • Biosolids proteins as potential labile C and N sources. Lerch et al. (1992) noted that extractable biosolids proteins were correlated to C mineralization (r= .95) but poorly correlated with N mineralization (r= .40). However, protein degradation products were correlated to N mineralization (r= .91). Overall results suggested that biosolids proteins are important sources of labile C that could enhance microbial activity as well as nutrient cycling and turnover and that protein degradation products are critical N sources for enhanced N cycling.
    • Biosolids protein extraction methodology and characterization. To simplify biosolids protein analyses, Lerch, Barbarick, et al. (1993) focused on several protein extraction methodologies. The authors ultimately recommended that Triton X-100 (a non-ionic detergent) be used as a routine extraction due to relatively short extraction times. After extraction, characterizing biosolids-borne proteins would help with understanding how soil microorganisms would utilize and degrade these compounds after biosolids land application. Lerch, Azari, et al. (1993) characterized biosolids-borne protein molecular weights, showing that the majority of these compounds were low molecular weight. These results suggested that (a) proteolysis occurred during wastewater treatment, and (b) biosolids proteins would be expected to degrade rapidly after land application due to microorganisms easily utilizing these low-molecular-weight compounds.
    • Phosphorus, Cu, Zn, Ni, and Mo concentrations in dryland wheat grain. Barbarick et al. (1995) modeled uptake of these elements over time, noting that (a) grain Mo concentrations did not change, potentially due to an antagonistic relationship with biosolids-borne Cu; (b) Ni concentrations were low yet linearly increasing with increasing biosolids applications over time (biosolids were likely acting as a labile Ni source); and (c) P, Cu, and Zn wheat grain accumulation followed a plateau model, meaning that winter wheat accumulated only certain quantities of P, Cu, and Zn and excluded the remaining. The latter findings helped dispute the “time bomb” theory, whereby biosolids-borne metals that would be associated with organic phases would be released over time, causing phytotoxicity within plants.
    • Predicting biosolids N mineralization to further support agronomic biosolids applications. After 10–12 yr of biosolids land application, Barbarick et al. (1996) focused on the 6.7 and 27 Mg biosolids ha−1 treatments in order to discern soil N distribution, N use efficiency, and mineralization rates. The authors noted that the 6.7 Mg ha−1 treatment (total N load applied was between 1,257 and 1,697 kg ha−1) resulted in an average of 54% of the added biosolids N present in the soil residual phase, 9% was removed by grain, and 38% was unaccounted for. However, at the 27 Mg ha−1 rate (total N load applied was between 5,028 and 6,788 kg ha−1), these average percentages shifted to 35, 2, and 63%, respectively. These data suggested greater N use efficiency at the 6.7 Mg ha−1 rate (i.e., the agronomic rate). The authors also noted that first-year net N mineralization rates were between 13 and 43% for the agronomic biosolids application rate; the greater biosolids rate resulted in greater first-year net N mineralization (41–67%). Findings suggested that the assumption of a 20% first-year N mineralization rate (i.e., the assumed rate in the mid-1990s) was underestimating N release from land-applied biosolids.
    • Correlations between cumulative biosolids-borne nutrient/trace element concentrations added and either ammonium bicarbonate-diethylenetriaminepentaacetic acid (AB-DTPA; soil extractable) or 4 M HNO3 (pseudo-total) soil metal concentrations. Barbarick et al. (1997) used 11 yr of data, noting that soil extractable elemental data provided better predictions (as compared with a pseudo-total soil metal extraction) for monitoring changes in Cd, Cu, Mo, Ni, P, Pb, and Zn in soils within biosolids land application programs. The AB-DTPA test may also help predict changes in grain Zn content within biosolids land application programs.
    • Extractable trace elements within the soil profile after 11 yr of biosolids land application. Barbarick et al. (1998) observed slight yet significant increases in Cd, Cr, Cu, Ni, Mo, Pb, and Zn concentrations within the zone of biosolids incorporation (0-to-20-cm depth). Most often, Zn was observed at greater concentrations below the 0-to-20-cm depth. Movement of Zn into the subsoil could help mitigate Zn deficiency symptoms in susceptible crops (Barbarick et al., 1998), as suggested above.
    • The influence of soil solution chemistry on metal mobility. Al-Wabel et al. (2002) noted that biosolids applications increased dissolved organic C and, concomitantly, Cu. However, the authors also noted that the fraction of Cu, as well as Zn and Pb, present as free metal ions or inorganic complexes was relatively low. This suggested that dissolved organic C may play a role in the mobility of metals. Furthermore, metal mobility was greatest immediately after biosolids land application and decreased over time (Al-Wabel et al., 2002).
    • Termination of excessive biosolids applications and effects on wheat yield, grain elemental characteristics, and soil extractable nutrient concentrations. Over time, Barbarick and Ippolito (2003) noticed that the 40 dry Mg ha−1 rate, applied biennially, was excessive. The authors terminated biosolids land application to these plots after they had received five biosolids applications and then monitored system recovery over time. The authors essentially found that, after 6 yr (i.e., three wheat–fallow cropping cycles), (a) soil extractable nutrients were similar to the controls, although soil NO3–N concentrations remained elevated, and (b) grain uptake rates of N, P, Zn, and Cu were similar to the control;
    • Biosolids-borne Ba in dryland wheat agroecosystems. Ippolito and Barbarick (2008) focused on biosolids-borne Ba fate within these dryland winter wheat agroecosystems. Since project inception, the authors noticed that increasing biosolids application rates increased total soil Ba concentrations, yet AB-DTPA extractable Ba concentrations tended to decrease. Up to 110 kg Ba ha−1 had been applied to plots receiving biosolids. Labile Ba content was found to decrease due to the formation of insoluble BaSO4 on the soil surface and potential BaCO3 formation in the subsurface. These Ba mineral precipitates led to soluble Ba concentrations one to three orders of magnitude lower then USEPA drinking water standards of 2 mg L−1 (USEPA, 2005).
    • The impact on soil P fate and transport. During the last year in which these plots were used, Ippolito et al. (2007) studied soil P accountability, fractionation, and potential environmental risk due to long-term, increasing biosolids application rates. Between 93 and 128% of the biosolids-borne P added was accounted for, with variations due to soil displacement due to tillage, removal of P with wheat grain, and soil sorption. Within the zone of biosolids incorporation (0–20 cm), Fe-P mineral phases dominated, likely due to increased Fe present in biosolids; Fe2(SO4)3 was added at the wastewater treatment facility to reduce H2S formation. Subsoil P was associated with Ca-P mineral phases. Taking these findings into consideration and using Colorado's Phosphorus Risk Index (Sharkoff, 2012), biosolids applications would not be impeded based on biosolids-borne P inputs.
    • The fate of biosolids-borne trace metals. Also during the last year that these plots were used, Ippolito and Barbarick (2008) assessed the distribution of Cd, Cr, Cu, Mo, Ni, Pb, and Zn in the 0-to-20-cm and 20-to-60-cm depths using pseudo-total metal digests (i.e., 4 M HNO3) and a sequential extraction technique designed by Sloan et al. (1997) for use in biosolids-amended soils. Based on the 4 M HNO3 digest, trace elements were concentrated in the soil surface, with no significant downward movement noted. The sequential extraction results showed that (a) Cd was present in relatively mobile forms; (b) Cr, Cu, Mo, Ni, Pb, and Zn were present in relatively immobile forms; and (c) increasing biosolids application rates increased Cu, Ni, and Pb present in relatively mobile phases yet also increased Cu, Mo, and Zn in relatively immobile phases. Overall findings suggested that long-term biosolids application rates at or above the suggested agronomic rate did not significantly contribute to downward metal transport, and thus groundwater contamination would not be affected by biosolids land application.
    Details are in the caption following the image
    Air-dried biosolids hand application to plots near Bennett, CO (courtesy of Jim Ippolito)

    2.6.2 North Bennett, CO, biosolids land application research program

    Newer plots near Bennett, CO, were established in 1993 and 1994 to more definitively identify a biosolids agronomic N rate for producing dryland winter wheat in Colorado. Prior to site establishment, the biosolids agronomic N rate (i.e., the rate needed to supply the N needs of dryland winter wheat) was hypothesized to be 6.7 Mg ha−1 (Lerch et al., 1990a). These newer plots bracketed that agronomic N rate, with plots receiving biosolids at rates of 0, 2.2, 4.4, 6.6, 8.8, and 11 dry Mg ha−1. These plots were paired with plots receiving increasing rates of inorganic N fertilizer (urea; 0, 22, 44, 66, 88, and 110 kg ha−1). All materials were fully incorporated into the top 20 cm of soil; all soils were classified as Aridic Argiustolls. Biosolids and inorganic N fertilizer were applied every other year from 1993 through 2016. Subsequent research on these plots focused on repeated biosolids application effects on:
    • Nitrogen fertilizer equivalency for dryland winter wheat. Determining the N fertilizer equivalency and understanding biosolids-borne N mineralization rates are both paramount for determining the biosolids agronomic rate for dryland winter wheat (Barbarick & Ippolito, 2000). Using crop N uptake data, Barbarick and Ippolito (2000) found that 1 dry Mg biosolids ha−1 delivered 8.2 kg N fertilizer ha−1. The authors also determined that first-year biosolids-borne N mineralization rates ranged from 25 to 32%; a subsequent study focused more closely on estimating first-year mineralization rates and N equivalency. Barbarick and Ippolito (2007) used data that encompassed both periods of greater and lower than average precipitation to determine first-year biosolids-borne N mineralization rates. The authors noted that biosolids N equivalency ranged between 8 and 9 kg N Mg−1 biosolids, similar to that found by Barbarick and Ippolito (2000). Based on winter wheat N requirements, biosolids agronomic application rates would be between 4.5 and 6.7 dry Mg ha−1. The authors also noted that, during periods of greater-than-average or lower-than-average precipitation, first-year N mineralization rates ranged from 25 to 32% and from 21 to 27%, respectively. Based on the findings of Barbarick and Ippolito (2000, 2007), the authors suggested that the first-year N mineralization rate should be estimated at 25%. This, in combination with the estimated biosolids N equivalency, has helped biosolids land applicators determine agronomic needs for crops. In addition, the information gleaned from this newer study, in conjunction with the original study, was used to help shape the State of Colorado biosolids plant available N and agronomic rate calculation (Colorado Department of Public Health & Environment, 2022).
    • Predicting availability of biosolids-borne soil nutrients over time. Barbarick and Ippolito (2008) used regression equations to better predict soil Zn, P, Fe, and Cu availability when using repeated biosolids applications over time. The authors found that planar regression models (r2 = .75–.92) predicted nutrient available within sites receiving multiple biosolids applications.
    • Predicting grain element concentrations under repeated biosolids land application scenarios. Barbarick and Ippolito (2009) used linear, quadratic, paraboloid, and exponential-rise-to-a-maximum equations for predicting winter wheat grain Ba, Cd, Cu, Mn, Mo, Ni, P, and Zn concentrations as a function of the number of biosolids applications over time. The authors showed that the paraboloid regression model was superior to other models in terms of estimating grain element uptake over time within a repeated biosolids land application program.
    • Long-term biosolids application effects on wheat yield, plant N uptake, and soil NO3–N alterations. Barbarick et al. (2010) used 15 yr of wheat yield and plant N uptake data, in combination with yearly fluctuations in soil NO3–N concentrations, to answer the following questions: (a) What is the relationship between cumulative grain yield and N removal and biosolids application rates and number of applications that have occurred? and (b) How are wheat grain yield and N uptake intertwined with residual soil NO3–N? Planar regression models showed that as yield or N uptake increased, residual soil NO3–N decreased. The modeling approach and findings could help biosolids land applicators make better decisions with respect to land application rates.
    • Elemental uptake coefficients for dryland winter wheat. The ratio of plant elemental concentration to quantity of element added (i.e., uptake coefficient) is used as a risk assessment tool by the USEPA (USEPA, 1993) for biosolids land application. Barbarick et al. (2014) determined uptake coefficients for Cu, Fe, Mo, Ni, P, and Zn after 10 applications of increasing biosolids rates at this site near Bennett, CO. Grain elemental concentrations followed no trend with biosolids application rates. Nutrient uptake coefficients followed an exponential decay model and subsequently did not provide useful information with respect to risk. However, Barbarick et al. (2014) did find useful information in the relationship between cumulative grain element removed compared with the cumulative element added, over time, with biosolids (r2 = .47–.99). The authors suggested that, in this system, the variability of uptake coefficients does not make its use as the best approach for assessing risk and that cumulative biosolids-borne element added to an ecosystem appears to be a better determinant of risk.
    • The extent of biosolids-borne plant nutrients to overall crop nutrient concentration, uptake, and ultimately off-site removal. Barbarick et al. (2016) used pathway analysis, in combination with multiple linear regression, to identify the (in)direct and total effects of repeated biosolids applications (over a 20-yr period), soil extractable elements, soil pH, and organic C content on removal of P, Zn, Cu, Fe, and Ni in wheat grain. The authors noted that (a) the number of biosolids applications had the greatest positive effect on grain nutrient removal, and (b) AB-DTPA–extractable Cu content was positively correlated with plant Cu uptake.
    • Long-term biosolids applications and their effects on soil health. During the last year of this project (2016) that was initially established in 1993 and 1994, Ippolito et al. (2021) collected soils from the top 20 cm and analyzed those soils for bulk density and water-stable aggregates (physical attributes); pH and EC (chemical attributes); extractable P and K (nutrient attributes); and soil organic C, potentially mineralizable N, microbial biomass C, and β-glucosidase activity (biological attributes). These soil characteristics were entered into the Soil Management Assessment Framework in order to discern long-term biosolids application effects on soil health. Increasing biosolids application rates increase soil chemical and biological health, leading to overall improvements in soil health. The results of Ippolito et al. (2021) attest to the long-term benefits of biosolids land application in dryland agroecosystems.

    2.6.3 Byers, CO, biosolids land application research program

    In 1999, CSU had the unique opportunity to assess agronomic biosolids land application (as compared with inorganic fertilizers) within no-till systems near Byers, CO; previous dryland research (above) was entirely focused on biosolids full incorporation into soils. When needed, based on soil test NO3–N and organic matter content, biosolids were surface applied (no incorporation) at an agronomic rate to wheat–fallow or wheat–corn–fallow test plots. Soils at this research site are classified as Ustollic Paleargids and Ustic Torriorthents. This research program, now in its 23rd year, has focused attention on:
    • Whether biosolids would produce the same grain yields and nutrient concentrations, and that biosolids-borne elements would not migrate substantially below the top 10 cm of soil, as compared with inorganic fertilizer applications. Using results from the first 10 yr of on-site research, Barbarick et al. (2012) noted that, as compared with inorganic fertilizer, (a) biosolids produced similar wheat and corn yields; (b) grain produced within biosolids plots contained lower Ba concentrations due to the formation of BaSO4 in soil, supporting findings of Ippolito and Barbarick (2008); (c) biosolids plots contained greater residual soil NO3–N with depth; (d) results were mixed with respect to some nutrients moving below the top 10 cm of soil; and (e) biosolids functioned as a Zn fertilizer to improve Zn accumulation in grain.
    • Earthworm (Aporrectodea caliginosa) survivability in biosolids-amended soils. McDaniel, Barbarick, et al. (2013) used soil (top 10 cm) from the Byers, CO, site that had received 11 Mg biosolids ha−1, adding earthworms to soils under varying degrees of drought stress (e.g., constant water content and 1-, 2-, or 3-wk cycles of water addition). The authors noted that (a) earthworm mass did not decrease with increased drought stress, (b) drought stress >2 wk resulted in greater estivation, and (c) drought stress >3 wk increased mortality rate to 14%.
    • Earthworm survivability and effect on nutrient availability. McDaniel, Stromberger, et al. (2013) collected the top 10 cm of unamended soil from the Byers, CO, site; returned it to the laboratory; and performed a study comparing unamended soil with soil to which the authors added 11 Mg biosolids ha−1. Earthworms were added to both soils, soil moisture content was maintained at 70% of field capacity, and soils were destructively sampled over 12 wk for survivability, microbial biomass C and N, and soil nutrient availability. The authors noted that (a) the majority (98%) of the earthworms survived, and survivability was not affected by biosolids; (b) biosolids did reduce earthworm weight gain; (c) microbial biomass C was not affected by earthworms, yet at times, a significant earthworm × biosolids interaction existed for microbial biomass N; furthermore, microbial biomass N tended to decrease over time, potentially due to increases in microbial turnover via earthworm activity (Hendrix et al., 1998); and (d) earthworms increased soil NH4–N but did not affect other quantified nutrients (e.g., plant available Ca, Cu, Fe, K, Mg, Mn, P, Zn).
    • Overall effects of biosolids in dryland wheat agroecosystems. Barbarick et al. (2017) performed meta-analyses on data collected from the Byers, CO, site (2001–2013) in conjunction with data collected from the “new” North Bennett, CO, site (1993–2013). The research premise was to understand whether biosolids application affects dryland wheat grain production and nutrient concentrations and available soil nutrients as compared with equivalent inorganic fertilizer applications. The highlight of this research was that biosolids increased available soil Zn concentrations from deficient to sufficient regardless of application method (i.e., fully incorporated or surface applied). Greater Zn availability led to greater grain Zn concentrations and resulted in wheat grain Zn biofortification. This is important to note because (a) winter wheat grain Zn deficiencies are prevalent within eastern Colorado (Miner et al., 2022) and likely across this region of the United States, (b) ∼31% of people globally are affected by Zn deficiencies (Caulfield & Black, 2004), and thus (c) biosolids may play a critical role in improving human health.
    • Research at this site will continue into the future, with scientists studying the long-term effects of biosolids application on soil health and the linkages between soil health, plant health, and human health (J. A. Ippolito, unpublished observation).


    Forty years of research within Colorado has proven the benefits of biosolids land application to a variety of ecosystems. In-state research has shown that sewage effluent could be land applied to benefit aboveground plant growth as well as to polish water prior to reaching receiving waters. Forest wildfire–affected ecosystems can benefit from biosolids applications, with application rates up to 80 Mg ha−1 leading to greater plant establishment, soil microbial activity, and nutrient turnover and reduced nutrient and heavy metal concentrations in runoff to below drinking water standards. Decadal-long observations in oil shale–mined lands showed that biosolids (up to 224 Mg ha−1) can positively influence microbial-mediated nutrient cycling and, in turn, aboveground plant community structure. Biosolids applications (up to 40 Mg ha−1) to desert shrubland ecosystems, dominated by Mo-containing shale deposits, can aid in reducing imbalances between Mo and Cu in soils and plants, helping to potentially offset molybdenosis when plants are consumed by ruminants. Biosolids and lime applications (both at 224 Mg ha−1) have been shown to improve long-term reclamation success on acid-generating, heavy metal–containing fluvial mine tailings. Three decades of grazing land research, focused on biosolids benefits to soils and aboveground plants, has shown that soil health and plant productivity can be improved to the greatest extent at biosolids application rates close to 10 Mg ha−1. Furthermore, four decades of dryland agroecosystem research has proven a multitude of benefits, including:
    • Disputing the “time bomb” theory by showing that plant metal uptake follows an exponential rise to a maximum (when mineralized, biosolids do not release exorbitant metal concentrations that could potentially become toxic to plants);
    • The long-term tracking of micronutrients and heavy metals in soils and plants has proven that soil concentrations will not lead to groundwater degradation and thatplants are safe for human consumption;
    • Biosolids provide Zn, helping to overcome soil deficiencies in calcareous soils (such as found in Colorado) and enhancing Zn biofortification in wheat grain;
    • Illustrating an economic return to producers via biosolids land application associated with increased wheat grain protein content, especially when protein premiums are paid to producers;
    • Identifying first-year N mineralization rates (25–32%);
    • Determining the biosolids N fertilizer equivalency of ∼8 kg N Mg−1.

    The last two points above are important to emphasize. When used in agroecosystems, biosolids are mineralized and thus can act as a slow-release fertilizer, supplying N needs during critical grain filling periods. Furthermore, given the ever-increasing cost of synthetic N fertilizers, biosolids can be used by producers as an alternative N fertilizer source. Ultimately, 45 yr of research in Colorado has proven that biosolids can enhance environmental quality, improve soil health, and produce healthy food products.


    This work was supported by the USDA National Institute of Food and Agriculture, Multi-State Hatch project COL00292D, accession 1020695.


      James A. Ippolito: Conceptualization; Data curation; Formal analysis; Funding acquisition; Investigation; Methodology; Project administration; Resources; Supervision; Validation; Visualization; Writing – original draft; Writing – review & editing. Ken A. Barbarick: Conceptualization; Data curation; Formal analysis; Funding acquisition; Investigation; Methodology; Project administration; Resources; Supervision; Validation; Visualization; Writing – original draft; Writing – review & editing.


      The authors declare no conflict of interest.