Biomass Accumulation, Contaminant Removal, and Settling Performance of Chlorella sp. in Unsterilized and Diluted Anaerobic Digestion Effluent

Abstract

Microalgae demonstrate significant efficacy in wastewater treatment. Anaerobic digestion effluent (ADE) is regarded as an underutilized resource, abundant in carbon, nitrogen, phosphorus, and other nutrients; however, the presence of inhibitory factors restricts microalgal growth, thereby preventing its direct treatment via microalgae. The purpose of this study was to dilute ADE using various dilution media and subsequently cultivate Chlorella sp. to identify optimal culture conditions that enhance microalgal biomass and water quality. The effects of various dilution conditions were assessed by evaluating the biomass, sedimentation properties, and nutrient removal efficiencies of microalgae. The results demonstrate that microalgal biomass increases as the dilution ratio increased. The microalgae biomass in the treatments diluted with simulated wastewater was significantly higher than that with deionized water, but their effluent quality failed to meet discharge standards. The treatment diluted with deionized water for 10 times exhibited abundant microbial biomass with strong antioxidant properties. The corresponding total phosphorus concentration in the effluent (6.96 mg/L) adhered to emission limits under the Livestock and Poultry Industry Pollutant Emission Standards (8 mg/L), while ammonia nitrogen concentration (90 mg/L) was near compliance (80 mg/L). The corresponding microbial biomass, with a sludge volume index (SVI₃₀) of 72.72 mL/g, can be recovered economically and efficiently by simple precipitation. Its high protein (52.07%) and carbohydrate (27.05%) content, coupled with low ash (10.75%), makes it a promising candidate for animal feed and fermentation. This study will aid in understanding microalgal growth in unsterilized ADE and establish a theoretical foundation for cost-effective ADE purification and microalgal biomass production.

1. Introduction

China generates approximately 4 billion tons of livestock manure annually, and inadequate management of this waste can lead to significant agricultural pollution [1]. In 2020, the Ministry of Agriculture and Rural Affairs of China, in collaboration with the Ministry of Ecology and Environment, issued a notice entitled “Further Clarifying the Requirements for Strengthening the Supervision of Breeding Pollution for the Return of Livestock and Poultry Manure to the Field.” This notice explicitly called for the comprehensive promotion of resource utilization of livestock and poultry breeding waste. Anaerobic digestion is recognized as a key strategy for converting livestock and poultry manure into usable resources. However, the large volumes of anaerobic digestion effluent (ADE) produced from the anaerobic digestion of livestock manure cannot be effectively managed through direct field application, seed soaking, or water–fertilizer integration [2]. Thus, there is an urgent need to develop a resource-efficient treatment method for ADE that not only ensures compliance with discharge standards but also facilitates the recovery and utilization of the nutrients contained in it.

Microalgae are characterized by a rapid growth rate, high photosynthetic efficiency, high value-added product contents, and the ability to fix CO₂. Despite their potential benefits, the cultivation of microalgae is currently hindered by relatively high costs, primarily attributed to the consumption of chemical fertilizers and freshwater, as well as the energy-intensive harvesting process [3]. Consequently, achieving large-scale cultivation of microalgae while controlling costs remains a significant challenge for the microalgae application industry. ADE is rich in essential nutrients for microalgae growth and reproduction, such as carbon, nitrogen, and phosphorus. Cultivating microalgae with ADE can recover these nutrients while also producing microalgal biomass, thereby reducing the costs of chemical fertilizer and freshwater and promoting the resource utilization of ADE. The complex composition of ADE, characterized by high concentrations of ammonia nitrogen, turbidity, heavy metals, and antibiotics, etc., presents significant challenges for microalgae cultivation, as these contaminants can inhibit their growth [4,5,6], leading to low efficiency in ADE purification. For example, excessive ammonia concentrations can significantly impair microalgae growth by disrupting photosynthetic processes. Specifically, ammonia can damage vital components such as photosystems I and II, electron transport chains, oxygen-evolving complexes (OECs), and dark respiration pathways [7]. Additionally, ADE contains substantial concentrations of heavy metals, such as copper and zinc, which can further inhibit microalgae growth [4]. Moreover, Chen et al. [8] reported that sulfonamides can increase the protein content of Chlorella sp., reduce soluble sugar levels, elevate oxidative stress, and alter the ultrastructure and DNA of microalgal cells. Given the dual objectives of microalgal cultivation and ADE purification, it is essential to pre-treating ADE to reduce the inhibitory effects of toxic and harmful substances on microalgal growth.

Common pretreatment technologies for ADE include the dilution method, autoclave sterilization, ozone treatment, and chemical precipitation [6,9,10,11,12], but each of these methods has certain limitations. Wang et al. [13] cultivated Micrococcus using sterilized ADE at various dilution ratios and found a linear negative correlation between the microalgal biomass productivity and the dilution ratio. Despite this, the ammonium removal rate in the ADE exceeded 90%, suggesting that sterilized ADE can be a suitable medium for microalgal cultivation. However, simple dilution reduces the concentration of nutrients essential for microalgal growth, thus limiting biomass accumulation, while sterilization increases cultivation costs, rendering it impractical for large-scale microalgal cultivation [14]. Tang et al. [15] diluted landfill leachate with recovered microalgae medium and used it for microalgae cultivation, resulting in increased biomass productivity and a lipid content of 27.6%. Therefore, wastewater with low turbidity and high nutrient content can be used to dilute the ADE, alleviating the issues associated with simple dilution. In addition to its high cost and it being impractical for large-scale microalgae cultivation, sterilization also leads to the loss of indigenous microorganisms. In contrast, unsterilized ADE retains indigenous bacteria that can work synergistically with microalgae to enhance pollutant removal through a bacteria–microalgae symbiosis [16]. Under co-culture conditions, with an inoculation ratio of Chlorella sp. to Shinella sp. YHB03 at 20:1, microalgal biomass productivity reached 80 mg/L/d, and the removal rates of ammonia nitrogen, total phosphorus, total organic carbon (TOC), and inorganic carbon (IC) were 72.2%, 48.3%, 65.0%, and 63.0%, respectively [17]. Microalgae are typically suspended in the culture medium, and current harvesting methods such as centrifugation and chemical flocculation are costly, posing a barrier to large-scale production. However, when cultured with unsterilized ADE, extracellular polymeric substances (EPSs) secreted by indigenous bacteria can form large aggregates to wrap microalgal cells together, reducing the costs associated with microalgae harvesting by rapid self-precipitation [5].

This study is designed to assess the effects of ADE on Chlorella sp. growth and changes in wastewater quality at various dilution ratios to determine the optimal dilution conditions. The diluted ADE was utilized to culture Chlorella sp. The growth of microalgae, water quality of ADE, size of symbiotic agglomerates, sedimentation properties, and biochemical composition of microalgae were thoroughly analyzed. This study provides theoretical support for the direct cultivation of microalgae following simple, efficient, low-cost, and easily scalable pretreatment of ADE.

2. Materials and Methods

2.1. Wastewater Characteristics and Microalgae Strain

Chemical agents were Analytical Reagents, purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). This experiment utilized two types of wastewaters: ADE and simulated wastewater. ADE was collected from a centralized anaerobic digestion plant in Xinyu City, Jiangxi Province, China. It was then allowed to stand at room temperature for 24 h and subsequently filtered with 8 layers of gauze for further experiment. Simulated wastewater is wastewater containing nutrients artificially prepared in the laboratory. The characteristics of ADE and simulated wastewater are presented in Tables S1 and S2.

Chlorella sp. FACHB-31 was purchased from the Freshwater Algae Culture Collection of the Institute of Hydrobiology (Wuhan, China) and cultured with modified BG-11 (Table S3) at a temperature of 26 ± 1 °C and continuous light (142 μmol·m⁻²·s⁻¹).

2.2. Experimental Design

In this study, deionized water and simulated wastewater were used to create a gradient dilution series of ADE to cultivate microalgae. The purpose of diluting ADE with synthetic wastewater is to replenish essential nutrients lost during the dilution process. Specifically, ADE was diluted to 6-, 8-, and 10-fold concentrations using deionized water and simulated wastewater. The groups diluted with deionized water (DW) were labeled as DW6 (deionized water dilution 6-fold), DW8 (deionized water dilution 8-fold), and DW10 (deionized water dilution 10-fold), while those diluted with simulated wastewater were labeled as SW6 (simulated wastewater dilution 6-fold), SW8 (simulated wastewater dilution 8-fold), and SW10 (simulated wastewater dilution 10-fold). Simulated wastewater (SW) served as the control group (CK). At the start of the experiment, microalgal cultures were inoculated at an initial concentration of 3 g/L (each liter of culture contains 3 g of moisture-laden microalgal cells, which correspond to a converted dry cell weight of 1.2 g/L) in 150 mL of culture medium within 500 mL conical flasks. The culture conditions were identical to those described in Section 2.1. All cultures were manually agitated three times daily and performed in triplicate for 10 days.

2.3. Analysis Methods

2.3.1. Analysis of Microalgal Growth and Photosynthetic Activity

Every two days, 3 mL of sample was collected and centrifuged at 8000 rpm for 5 min. The microalgae particles were resuspended in 3 mL of deionized water. The samples were then diluted 6-fold with deionized water, and the absorbance at 680 nm was measured using a UV–vis spectrophotometer (UV-9000, Metash, Shanghai, China). The dry weight of the microalgae was calculated from a standard curve correlating absorbance values with microalgae dry weight (Texts S1 and S2).

The Fv/Fm ratio, which indicates the maximum quantum efficiency of photosystem II (PS II) photochemistry under dark conditions, was used to determine the photosynthetic activity of the microalgae and measured after incubating the samples in the dark for 15 min using a Water PAM fluorometer (Heinz Walz GmbH, Effeltrich, Germany).

2.3.2. Analysis of Pigment in the Microbes

Every two days, 1.5 mL of the sample was collected and centrifuged at 10,000 rpm for 5 min. The supernatant was discarded, and the pellet was resuspended in 1.5 mL of methanol. The suspension was incubated in darkness at 45 °C in a constant-temperature water bath for 24 h. After incubation, the sample was centrifuged again at 10,000 rpm for 5 min. One milliliter of the supernatant was then taken and diluted threefold with methanol. Absorbance values were measured using an ultraviolet–visible spectrophotometer (UV-9000, Metash, China) at 470 nm (A₄₇₀), 652.4 nm (A_652.4), and 665.2 nm (A_665.2). The microalgal pigment concentration was calculated using the method described by Lichtenthaler H K et al. [18].

$$\mathrm{Chla} (\mathrm{mg}/\mathrm{L})=16.72{\mathrm{A}}_{655.2}-9.16{\mathrm{A}}_{652.4}$$

(1)

$$\mathrm{Chlb} (\mathrm{mg}/\mathrm{L})=34.09{\mathrm{A}}_{652.4}-15.28{\mathrm{A}}_{665.2}$$

(2)

$$\mathrm{Carotenoids} (\mathrm{mg}/\mathrm{L})={\displaystyle \frac{1000{\mathrm{A}}_{470}-1.63\mathrm{Chla}-104.96\mathrm{Chlb}}{221}}$$

(3)

2.3.3. Analysis of Wastewater Quality

Wastewater quality was assessed using the method outlined by Gu et al. [19]. Prior to analysis, samples were filtered through a 0.45 μm aqueous microporous filter membrane. For ammonia nitrogen (NH₄⁺–N) and total phosphorus (TP) analysis, samples were properly diluted with deionized water (DW) and analyzed using a water quality analyzer (5B-6C, Lianhua, Beijing, China). For total organic carbon (TOC) and inorganic carbon (IC) analysis, samples were properly diluted with DW and analyzed using a TOC/TN analyzer (Multi N/C 3100, Analytik Jena AG, Jena, Germany). The pH of the wastewater samples was measured with a pH electrode (PH-3G, REX, Shanghai, China). The specific methods for the determination of heavy metals are presented in Text S3.

2.3.4. Analysis of Oxidative Stress in the Microbes

After 10 days of culture, 3 mL samples were centrifuged, resuspended in 0.1 M phosphate buffer solution (PBS, pH 7.2–7.4, Solarbio), and disrupted in an ice bath to determine the superoxide dismutase (SOD) activity and soluble protein indexes of the microbes. One milliliter of the sample was centrifuged and resuspended in PBS for reactive oxygen species (ROS) determination, 0.1 mL of DCFH-DA (Catalog No. 4091-99-0, Solarbio, Beijing, China) probe was added, and the fluorescence intensity was measured by a fluorescence spectrophotometer (F-380, LabTech, Dongguan, China) after incubation. After staining the microalgae cells with live and dead dyes (L7012, Thermo, Waltham, MA, USA), visible light and green and red fluorescence images were taken with a fluorescence microscope (EVOS M7000, Thermo, USA). Refer to Text S4 for detailed parameters and operational methods pertaining to oxidative stress parameters.

2.3.5. Analysis of Microbes Settlement Performance

After 10 days of cultivation, 50 mL of the culture was transferred into a 50 mL colorimetric tube to assess the settling performance. The culture was left to settle for 30 min, after which the supernatant was carefully removed. The volume (V) of the settled biomass at the bottom was recorded, followed by drying and weighing to determine the mass (M) of the microbes. The sludge volume index (SVI) after 30 min (SVI₃₀) was then calculated using the following formula:

$$\mathrm{SV}{\mathrm{I}}_{30}={\displaystyle \frac{\mathrm{V}}{\mathrm{M}}}$$

(4)

where V is the volume (mL) of the settled biomass, and M is the weight (g) of the dried biomass after 30 min of settling.

The particle size of the flocs in the culture was measured using a laser particle size analyzer (BT-9300HT, Dandong Bettersize Instruments Ltd., Dandong, Liaoning, China) with deionized water as the medium. The shading range was set between 5% and 15%. Prior to measurement, the flocs in the wastewater sample were rinsed three times with deionized water to perform background correction. The stirrer was kept in operation to ensure the uniform distribution of the sample in the medium. To remove air bubbles, the ultrasonic timer was activated for 3 min. After this, the sample was added for analysis. The “single-time” measurement mode was selected, and the results were exported to an Excel file. The Excel file presents the values of D10, D50, and D90, which represent the cumulative percentage of particles less than specific sizes. For example, D50 for 130 μm implies that the proportion of particles less than 130 μm is 50%.

2.3.6. Analysis of Biomass Biochemical Composition

The biochemical composition of the harvested biomass was determined using the method described by Zhang et al. [20]. In brief, the crude protein content was measured using an automatic Kjeldahl nitrogen analyzer (K9860, Shandong Haineng Instrument, Dezhou, China). The crude lipid content of the microalgae was determined via Soxhlet extraction (n-hexane as the solvent, extraction temperature of 90 °C, and extraction time of 3 h). Ash content was measured after calcination in a muffle furnace (YX-1000XB, Yaoxing high-temperature kiln, Luoyang, China) at 550 ± 5 °C for 4 h. Carbohydrate content was calculated by subtracting the crude protein, crude lipid, and ash content from the total biomass.

2.4. Statistical Analysis

All experiments were conducted in triplicate, and the results are presented as mean ± standard deviation. The figures were drawn using Origin 2023 (OriginLab, Northampton, MA, USA). Statistical analysis was performed by one-way analysis of variance (ANOVA) using SPSS 16.0 (SPSS version 23.0, SPSS Inc., Chicago, IL, USA). Post hoc comparisons were made using either the least significant difference (LSD) test or a two-tailed t-test. A significance level of p < 0.05 was considered statistically significant.

3. Results and Discussion

3.1. Biomass Production and Microbial Activity

As shown in Figure 1a, the growth of Chlorella sp. was relatively slow when diluted with DW. Although Chlorella sp. initially exhibited some growth, there was no significant difference in biomass accumulation during the early stages. After day 4, the biomass accumulation increased with higher dilution ratios. As demonstrated by He et al. [21], the turbidity of the culture medium inhibits the growth of Chlorella sp. In the experiment, although ADE was diluted, the turbidity remained relatively high, which resulted in slower growth of Chlorella sp. at lower dilution levels. While the chlorophyll-a (Chl-a) content does not directly correlate with microbes biomass [22], it serves as an indicator of microalgal growth. In Figure 1c, the Chl-a content in the DW6 and DW8 groups increased from day 0 to day 2 and then stabilized, whereas in the DW10 group, the Chl-a content continued to rise. This suggests that Chlorella sp. experienced a lag phase in the early stage, leading to slow growth. In the DW10 group, interactions between Chlorella sp. and bacteria facilitated rapid growth. During autotrophic growth, microalgae do not directly utilize atmospheric CO₂ but rather use dissolved inorganic carbon species such as CO₃²⁻ or HCO₃⁻ for photosynthesis. This process releases OH⁻ into the culture medium, resulting in an increase in pH [23]. Meanwhile, bacterial metabolism of organic matter releases H⁺, lowering the pH, which eventually stabilizes the pH around 6.5 (Figure 1d).

In the treatments diluted with SW, Chlorella sp. biomass showed rapid accumulation during the first 6 days, followed by a slight decrease. Growth was most pronounced in the SW10 group. In the study by Huang et al. [24], it was demonstrated that the relative growth rate of microalgae in the logarithmic stage increased with the addition of 1 g/L acetate, as shown in the study by Castillo et al. [25], where the biomass of microalgae in mixed nutrient culture was higher than that of autotrophs and heterotrophs alone. Our study corroborated these findings. The primary reason for the increase in microalgae biomass concentration is the increased dilution, decreased turbidity, improved light penetration into the medium, and more adequate mixed nutrients for microalgae.

From day 0 to day 2, Chl-a content and pH values both increased, then fluctuated before declining. In the SW6 group, the pH decline was delayed, but the biomass concentration continued to rise. This suggests that when simple organic matter is abundant, both microalgae and bacteria rapidly grow. With increased dilution, the concentration of native bacteria and medium turbidity decreased, allowing for better microalgal growth. Compared with the control group, Chlorella sp. growth was significantly inhibited in the SW6 and SW8 groups. The delayed pH decline in the SW6 group may be due to slower microalgal growth, which limited the nutrients available for bacterial growth and reproduction. Overall, Chlorella sp. growth was better with SW dilution treatment groups than with DW dilution treatment groups, likely due to sodium acetate in SW dilution treatment groups that promoted the cell growth. As an intermediate product of the tricarboxylic acid (TCA) cycle, acetate can enhance nutrient absorption in both microalgae and bacteria [26], resulting in faster and greater biomass accumulation.

Microalgae can grow under both photoautotrophic and heterotrophic mode. Under photoautotrophic mode, microalgae produce their own organic matter through photosynthesis, whereas under heterotrophic mode, they synthesize biomass by absorbing organic matter from the medium. Temperature, light, and other factors can affect the photosynthesis of microalgae [27]. As shown in Figure 1b, the Fv/Fm of the experimental group diluted with DW groups was consistently above 0.6, indicating improved photosynthetic activity. Photosynthetic activity increased with the dilution ratio, with the DW10 group exhibiting the highest activity.

The content of Chl-a (mg/g) in Chlorella sp. fluctuated, but its trend mirrored that of photosynthetic activity (Figure 1e). At lower dilution ratios, the medium’s turbidity was relatively high, requiring light to penetrate the turbid ADE, which can result in larger Chl-a content in Chlorella sp. cells. Increasing the dilution ratio reduced the medium’s turbidity, enhancing photosynthetic activity and subsequently increasing Chl-a content. In the treatments diluted with SW, the photosynthetic activity was comparable with that of deionized water dilution, with Chl-a content (Figure 1e) rising with dilution ratios. However, total Chl-a content decreased over time, likely due to biomass accumulation outpacing chlorophyll synthesis. The SW6 group displayed lower photosynthetic activity compared with other groups. In the SW6 group, the concentration of inhibitors was the highest, resulting in the strongest stress, which affected photosynthesis and Chl-a synthesis. At the same dilution ratios, the photosynthesis of DW treatment groups is better than that of SW treatment groups, consistent with the results of Kong et al. [28].

Organic matter in the culture system affects Chlorella sp. photosynthesis. Under the mixotrophic mode, microalgae use CO₂ to absorb light energy for photosynthesis and obtain carbon and energy from organic matter to synthesize biomass. The control group exhibited higher photosynthetic activity than all experimental groups, further indicating that inhibitory factors in ADE adversely affected Chlorella sp. photosynthetic activity. As described by Mao et al. [29], the combination of artificial photosynthesis and heterotrophic microalgae demonstrates an efficient capacity to capture carbon and produce high-value products.

Carotenoids are ubiquitous pigments in photosynthetic organisms, playing diverse biological roles [30]. As depicted in Figure 1f, carotenoid concentrations in the DW treatment groups exhibited a gradual increase, with higher levels observed in the DW8 and DW10 groups compared with the DW6 group. The Chl a/b ratio initially surged but subsequently declined rapidly in the DW6 group. Conversely, the DW8 and DW10 groups initially showed a rapid increase followed by stabilization, with the ratio also increasing with a higher dilution ratio. Notably, under the same dilution ratios, the carotenoid content was consistently higher in the DW treatment groups than in the SW treatment groups. Carotenoids are known to enhance light absorption and protect photosystems from oxidative damage caused by reactive oxygen species during photosynthesis [31]. Therefore, the observed increase in carotenoid content is a response to mitigate oxidative stress in the microalgal culture [32].

The Chl a/b ratio in Chlorella sp. serves as an indicator of chloroplast light-trapping sensitivity. Higher ratios correlate with greater light sensitivity [17] and enhanced photosynthetic electron transport [33]. In the treatments diluted with DW (Figure 1g), the DW10 group exhibited the highest Chl a/b ratio. By regulating gene expression to reduce Chl b accumulation, the Chl a/b ratio can be increased, leading to higher biomass yields [34]. Similar results were observed in this study when an inhibitor in the culture medium was reduced through dilution, confirming the feasibility of this approach. In the treatments diluted with SW, the Chl a/b ratio initially increased rapidly but then decreased in the SW6 group. In contrast, the SW8 and SW10 groups showed a more stable increase. The differences in Chl a/b ratios between DW treatment groups and SW treatment groups can be attributed to the presence of easily available organic substances in wastewater, which may reduce microalgal light sensitivity and consequently lower the Chl a/b ratio.

3.2. Oxidative Stress Analysis

As shown in Figure 2a, ROS levels significantly decreased with an increasing dilution rate using DW. The treatment of DW10 exhibited the lowest ROS levels, and SOD activity followed a similar trend, indicating that dilution effectively reduced oxidative damage caused by growth-inhibiting substances in ADE. Although the DW10 group had the lowest SOD content (Figure 2b), it exhibited better growth than the DW6 group. This might be attributed to insufficient SOD production in the DW6 group to counteract ROS-induced oxidative damage. The balance between oxidative stress and resistance to oxidative stress was disrupted, consistent with the findings presented by Zhu et al. [35]. The observed decrease in pigment and photosynthetic activities in the DW6 group negatively impacted biomass accumulation. In the treatments diluted with SW (Figure 2a), ROS levels decreased with increased dilution rates. Similar to DW treatment groups, the lowest ROS levels were observed in the SW10 group. The SW6 group showed higher ROS levels, possibly due to the high concentration of inhibitory substances that caused toxicity to Chlorella sp. While Chlorella sp. growth was inhibited in the SW6 group, the SW10 group exhibited lower ROS levels. This might be because the addition of small molecular organic matter (sodium acetate) supported Chlorella sp. growth and enhanced its resistance to oxidative damage. This is also why the SW dilution groups had lower ROS levels compared with DW dilution groups under the same dilution ratio.

While ADE contains heavy metals that can hinder microalgal growth, soluble proteins within the microalgal cells can bind to these contaminants, forming chelates that neutralize their toxic effects [36]. Regardless of whether DW or SW was used to dilute ADE (Figure 2c), soluble protein content exhibited a similar decreasing trend with an increasing dilution rate, showing significant differences (p < 0.05). Under the same dilution ratios, SW treatment groups had higher soluble protein content than DW treatment groups, and all groups had higher soluble protein content than the control group (p < 0.05). Since SW contains small organic molecules that promote microalgal growth, the SW treatment groups exhibited faster microalgal growth and increased soluble protein content, enhancing resistance to the toxicity of heavy metals in ADE. This experiment demonstrated that increasing the dilution ratio reduced the concentration of inhibitors and minimized damage to microalgae. Both the DW and SW treatment groups displayed lower oxidative stress levels and greater resistance.

3.3. Nutrient Removal in Wastewater

In the treatments diluted with DW (Figure 3a), it can be observed that the TOC content in the culture medium remained at the initial level of the experiment no matter which dilution ratio was applied, and the TOC content did not decrease; the TOC content even increased at a dilution rate of 6 and 10, indicating that Chlorella sp. and indigenous microbes cannot utilize TOC in ADE. This occurs because, after anaerobic digestion, most of the organic carbon in livestock manure is converted into biogas in the form of methane, leaving behind organic matter that is difficult for both Chlorella sp. and other organisms to utilize [37]. In future studies, bacteria or alternative treatment methods can be employed to decompose this challenging organic matter and convert it into forms that are usable by microalgae, significantly reducing the costs associated with ADE and microalgae cultivation. The IC content in all treatment rapidly decreased and then tended to be stable (Figure 3b), indicating the utilization of IC by microalgae during photosynthesis. The low IC content in the culture medium of DW10 and SW10 led to the corresponding lower IC removal efficiency. TOC rapidly decreased in all treatment groups diluted with SW on 0–2 days (Figure 3a) and then tended to be stable, while IC changed little on 0–2 days (Figure 3b), indicating that Chlorella sp. experienced mixotrophic growth at this stage, with heterotrophic dominating and autotrophic playing a supplementary role [38].

At the same dilution ratios (Figure 3a), TOC contents treated with a different dilution medium were similar after 2 days of culture, which further indicated that TOC contained in the medium cannot be used by Chlorella sp. and indigenous microbes. In this experiment, SW was used to dilute ADE, sodium acetate was introduced into the culture medium, and Chlorella sp. can perform heterotrophic growth and accelerate the accumulation of biomass. In order to be more consistent with the actual production, the culture medium was not sterilized, and the initial concentration of indigenous microbes in the culture medium was decreased with an increasing dilution rate. In experiments, changes in the proportion of bacteria and microalgae may not have contributed to enhanced wastewater treatment, which is inconsistent with previous studies on the ability of bacteria to enhance the purification of wastewater. Studies show that coordinated and balanced microalgae and bacterial population structures are essential to optimize wastewater treatment performance [39].

Ammonia nitrogen (NH₄⁺–N) constitutes over 90% of the total nitrogen in ADE. Chlorella sp. can tolerate NH₄⁺–N concentrations up to 100 mg/L, with low concentrations promoting growth and high concentrations inhibiting it [40]. In this study (Figure 3c), the highest NH₄⁺–N concentration in the medium was 532.44 mg/L (SW6), while the lowest was 213.48 mg/L (DW10), potentially inhibiting Chlorella sp. growth. As shown in Figure 3c, NH₄⁺–N levels in the DW treatment groups slowly declined initially, then rapidly from days 2 to 6, and subsequently slowed down, maintaining a downward trend throughout the experiment. In the treatments diluted with SW, NH₄⁺–N exhibited a rapid initial decline followed by a slower decrease, also ending with a downward trend. Microalgae can absorb and assimilate NH₄⁺–N into amino acids [41], as it is the optimal nitrogen source for their growth and can be directly utilized without further processing [42]. Consequently, the reduction in NH₄⁺–N concentration observed in this experiment may be directly absorbed and utilized by the microalgae.

Under the same dilution ratio (Figure 3c), the DW treatment group exhibited a lower NH₄⁺–N removal rate compared with the SW treatment group, likely due to the rapid growth rate of Chlorella sp. in the SW treatment group, which utilizes NH₄⁺–N more effectively. These results indicate that the NH₄⁺–N concentration in unsterilized ADE is adequate to support Chlorella sp. growth. Wastewater with low NH₄⁺–N can be used to dilute ADE. In the DW10 group, the concentration of NH₄⁺–N was expected to meet the Livestock and Poultry Farming Pollutant Discharge Standard (GB 18596-2001 [43]). Chlorella sp. utilized IC in the culture medium for photosynthesis, which led to an increase in pH. When the pH exceeds 8, NH₄⁺–N becomes free ammonia [44]. Free ammonia is unstable in the culture medium and prone to volatilization. Consequently, NH₄⁺–N in ADE can be removed through ammonia volatilization and Chlorella sp. ammonia assimilation.

Chlorella sp. requires phosphorus sources for growth and reproduction, as phosphorus is essential for synthesizing nucleic acids, phospholipids, ATP, and other compounds [45]. Bacteria also necessitate phosphorus for growth and reproduction. As depicted in Figure 3d, the TP content in the DW treatment groups initially increased but then gradually decreased. However, the final TP content was higher than the initial level. The DW8 group exhibited the most significant phosphorus content change, while the DW10 group reached a final phosphorus content of 6.96 mg/L, complying with the Livestock and Poultry Farming Pollutant Discharge Standard (GB 18596-2001).

The increase in phosphorus content might be attributed to the oxygen produced by Chlorella sp. during photosynthesis, leading to the death of anaerobic bacteria and subsequent phosphorus release into the culture medium. In the treatments diluted with SW, TP content remained relatively stable, with a slight decrease during the process. Regardless of the dilution methods, the phosphorus source content in the culture medium did not significantly decrease. The results indicated that the growth and reproduction of microalgae can be sustained within a phosphorus concentration range of 0.03% to 0.06% in the medium [46,47]. In this study, the TP content was adequate to support the growth and reproduction of microalgae. This suggests that the unsterilized ADE itself contains sufficient phosphorus to support Chlorella sp. growth and reproduction without additional supplementation. The DW10 group, meeting the Livestock and Poultry Farming Pollutant Discharge Standard (Second Exposure Draft for GB, 18596-2001), offers a promising foundation for further wastewater treatment efforts.

3.4. Settling Performance

Traditional microalgae harvesting methods based on gravity sedimentation are inefficient due to the small size (5–20 μm) of microalgae cells [48]. This has led to the adoption of centrifugation, flocculation, and filtration techniques [49]. Flocculation, achieved by adding inorganic or organic matter to neutralize the negative surface charge of microalgae, facilitates solid–liquid separation [50]. However, these additives can be difficult to remove from the harvested microalgae biomass, limiting its subsequent applications [51]. Additionally, these methods increase harvesting costs. Zheng et al. [52] successfully used self-flocculating microalgae to treat aquaculture wastewater, achieving efficient nitrogen and phosphorus removal. In this experiment, the unsterilized culture medium contained numerous bacteria and particulate matter, which can potentially act as flocculants. The observed good settling performance of the biomass in the experimental groups suggests that their settling behavior can be a valuable theoretical basis for solid–liquid separation after ADE treatment, significantly reducing harvesting costs of biomass.

In the sedimentation images (Figure 4e), it can be seen that most of the biomass settled down within 5 min, and in the following 1 h and 6 h sedimentation, although the culture medium was stratified, the sedimentation efficiency was similar to that within 30 min. Therefore, the sedimentation efficiency of 30 min (SVI₃₀) was determined. The SVI is an important indicator of settlement performance, and it is generally believed that the value of 50–120 mL/g has better settlement performance. In the treatments diluted with DW (Figure 4a), SVI₃₀ values were all between 50 and 100 mL/g, among which DW10 group had the highest SVI₃₀ value, which was significantly different from DW6 and DW8 groups (p < 0.05), indicating the good sedimentation performance after dilution. In the treatment groups diluted by SW (Figure 4a), the SVI₃₀ of the SW8 group was the highest, but the value of all treatments was lower than 50 mL/g, indicating that the settlement performance was poor. As can be seen from the sedimentation photos, although there is precipitation at the bottom, the upper liquid is green, indicating that most of the microalgae cells are still suspended in the medium.

Laser particle size analysis was used to measure the particle size of the flocs, providing a more direct reflection of sedimentation performance. The D10 values reflect the size of the smaller particles; smaller values indicate the presence of more fine particles in the system, potentially hindering the settling performance of the overall system. The D50 values reflect the average particle size of the system, commonly employed to characterize the overall particle size distribution of the sample. In contrast, the D90 values indicate the presence of larger particles and facilitate faster sedimentation [53]. As shown in Figure 4b–d, in the treatments diluted with DW, D10, D50, and D90 values all increased with a higher dilution ratio. The DW10 group exhibited the largest values (37.92, 130.00, and 288.13 μm, respectively), suggesting that higher dilution ratios facilitate the formation of aggregates between microbes. The larger particle sizes in the DW10 group might be due to a decrease in bacterial concentration and tighter binding between microbes. In the treatments diluted with SW (Figure 4b–d), the SW8 treatment group had significantly higher D10, D50, and D90 values (7.94, 111.73, and 278.67 μm, respectively) compared with the SW6 and SW10 groups (p < 0.05). Under the same dilution ratio, DW treatment groups had larger particle sizes than SW treatment groups. This can be due to increased available nutrients in SW-diluted groups, leading to faster microalgal cell growth and higher cell concentrations. As observed in the visible light images, although large particle aggregates were present in the treatment diluted with SW, most microalgal cells remained suspended. Compared with the control group, the significantly increased D10, D50, and D90 values in the treatment groups confirmed that substances (bacteria or particles) in ADE can form large aggregates with microalgae, accelerating sedimentation, similar to the finding of Kosar et al. [54].

3.5. Biochemical Composition

When diluted with DW (

Table 1. Biochemical composition of the harvested biomass.

Biochemical Composition	CK	DW6	SW6	DW8	SW8	DW10	SW10
Crude lipid (%)	23.6 ± 1.24 a	17.9 ± 0.96 b	22.0 ± 1.76 a	7.5 ± 0.63 d	23.3 ± 0.21 a	10.1 ± 0.62 cd	13.3 ± 0.89 c
Crude protein (%)	54.60 ± 0.09 a	50.08 ± 1.02 b	52.21 ± 1.86 ab	50.68 ± 1.07 b	53.00 ± 0.47 ab	52.07 ± 0.56 ab	51.71 ± 0.88 ab
Crude ash (%)	7.37 ± 0.44 d	13.69 ± 0.21 b	16.32 ± 0.08 a	11.09 ± 0.68 c	16.78 ± 0.37 a	10.75 ± 0.05 c	10.26 ± 0.38 cd
Carbohydrate (%)	14.61 ± 1.10 e	18.29 ± 0.26 d	9.47 ± 0.18 f	30.72 ± 0.25 a	6.88 ± 0.11 g	27.04 ± 0.11 b	24.76 ± 0.38 c

Note: Superscripts with the same letter in the same line indicate no significant difference, while different letter superscripts indicate a significant difference.

), the lipid content decreases as the dilution rate increases. The lipid content of the DW6 group was significantly higher than that of the DW8 and DW10 groups (p < 0.05). Typically, nitrogen limitation results in increased levels of microalgal lipids [55]; however, the experiment observed a downward trend, which is inconsistent with previous studies. In the experiment, although diluted, the concentration of ammonia nitrogen remains sufficient for the growth of microalgae, indicating that nitrogen restriction is unlikely to be a factor [56]. The protein content slightly increased with the dilution ratio but showed no significant differences among the three treatment groups (p > 0.05). The DW10 group had slightly higher protein content than the other two groups, while its ash content was significantly lower (p < 0.05). The carbohydrate content significantly varied among the three treatment groups (p < 0.05); the DW8 group had the highest carbohydrate content, followed by DW10 and DW6.

In the treatments diluted with SW (Table 1), the lipid content also decreased with an increasing dilution ratio. The SW10 group had significantly lower lipid content than the other two groups (p < 0.05), while the protein content did not significantly differ (p > 0.05). The SW10 group also had significantly lower ash content. Studies have shown that nutrients can promote Chlorella sp. lipid synthesis [28]. The presence of sodium acetate in SW might have contributed to the higher lipid content in SW-treated groups compared with DW-treated groups. The control group had higher lipid and protein content than the experimental groups, indicating that substances in ADE influenced microalgal lipid and protein accumulation. The control group also had the lowest ash content, suggesting the presence of sludge particles that facilitated sedimentation but increased ash content of the biomass. A high lipid content in microalgae is beneficial for biodiesel production, while a high protein content is suitable for animal feed applications. A low ash content is desirable for raw materials of feed, and a high carbohydrate content is advantageous for anaerobic digestion. Chlorella sp. cultured with DW-diluted ADE exhibits high carbohydrate and protein content, making it suitable as a substitute for anaerobic digestion and feed. Chlorella sp. cultured with SW-diluted ADE has a high lipid content, making it suitable for biodiesel production.

4. Conclusions

For efficient and low-cost production of microalgal biomass and purification of wastewater, anaerobic digestion effluent (ADE) was diluted with a different medium at various dilution ratios. The sedimentation performance of microalgae was subsequently evaluated. While the simulated wastewater dilution treatment group yielded higher biomass than the deionized water dilution treatment group, the introduction of numerous nutrients hindered compliance with discharge standards. The 10 times dilution treatment group of deionized water has high biomass yield and strong oxidative stress tolerance. Total phosphorus in wastewater meets the Standards for Emissions of Contaminants from Livestock and Poultry Farming (GB, Draft Second Exposure in 18596-2001), and ammonia nitrogen is also expected to meet the standards. The sedimentation experiment demonstrated excellent settling performance, with most microalgae cells depositing at the bottom, facilitating cost-effective biomass recovery. The harvested microalgae biomass exhibited high protein and carbohydrate content and low ash content, making it suitable as raw materials of feed and fermentation.