INTRODUCTION
The acceleration of urbanization has made pollution problems increasingly prominent, among which urban sewage has become a major issue. Wastewater generated from industrial production and daily life not only seriously pollutes rivers and water sources but also poses a great threat to human health and the ecological environment. Therefore, improving the quality of urban sewage treatment to ensure the healthy development of cities has become an urgent problem to be solved.
Traditional chemical treatment methods, if operated improperly, can easily cause secondary pollution and change the natural structure of unpolluted water bodies. In contrast, bioaugmentation technology based on biofilm, as an emerging means of urban sewage treatment, shows unique advantages. This technology can not only avoid secondary pollution but also has the characteristics of simple operation and real-time treatment. More importantly, bioaugmentation technology is less affected by pollutant concentration and salinity, maintaining high activity even under adverse conditions, and does not cause pollutant transfer when treating toxic wastewater.
However, the actual treatment effect of biofilm is influenced by many factors, among which the type of carrier is one of the key factors. Different carriers have different effects on biofilm growth and performance. Therefore, selecting appropriate carriers is crucial for improving the effectiveness of bioaugmentation technology in water pollution treatment. To this end, this paper conducts an in-depth analysis and research on the bioaugmentation treatment method for urban water pollution in environmental protection engineering through experiments, aiming to find the best carrier type and operating conditions, and to provide more effective technical support for urban water pollution treatment.
1 EXPERIMENTAL MATERIALS AND METHODS
1.1 Experimental Setup
The experiment used a plexiglass biofilm reactor with dimensions of 20 cm x 50 cm and an effective volume of 6.85 L. The aeration device was an oxygen-increasing pump, whose main function was to supply oxygen to microorganisms and ensure sufficient suspension of the suspended carriers. During the 3 h intermittent aeration period, a stirrer was used to fully mix the solid-liquid (microbial carriers and sewage). The reaction temperature was controlled at 28 ± 1°C using a temperature control device. The overall experimental setup is shown in Figure 1.
As shown in Figure 1, reactors (a) and (b) were inoculated with sediment, with carriers being suspended carriers and elastic carriers, respectively; reactors (c) and (d) were inoculated with activated sludge, with carriers being suspended carriers and elastic carriers, respectively. The operating parameters of the four reactors (a)~(d) were basically the same. The operation mode was 3 h + 3 h intermittent aeration with gradually increasing aeration intensity and intermittent sludge discharge. The aeration rate was 3 L/min for days 1–2, 2 L/min for days 3–9, and 3 L/min after day 9. The biofilm attachment method was complete discharge after 2 days of intermittent operation.
1.2 Experimental Materials
1.2.1 Configuration of Microbial Carriers
Elastic carrier configuration: The experiment used a 4 mm diameter corrosion-resistant central rope串联 about 600 filamentous substances made of high-density polyethylene as the elastic carrier, which was fixed at the top of the biofilm reactor and hung naturally. The filling height of this carrier was 26 mm, filament diameter 10 mm, mass 0.978 g/cm³, specific surface area 500 m²/cm⁻³, porosity greater than 95%, and filling ratio 60%.
Suspended carrier configuration: The suspended carriers were granular substances made of polyvinyl chloride, about 260 particles, which were in a fluidized state during aeration. The carrier diameter was 150 mm, particle diameter 0.5 mm, mass 100 g/cm³, specific surface area 311 m²/cm⁻³, porosity greater than 99%, and filling ratio 60%.
1.2.2 Preparation of Experimental Water
The experimental water was prepared based on the water quality of the Mihe River in Weifang City, Shandong Province. Common pollutants in Mihe River water, such as KNO₃, NH₄Cl, C₆H₁₂O₆, and NaNO₂, were added. The water temperature was controlled at 7.8°C–31.8°C, pH at 7.01–8.66, DO at 6.46–11.35 mg/L, NH₄⁺-N at 0.35–6.51 mg/L, and COD at 102–273 mg/L. Trace elements such as Cu²⁺, Mg²⁺, and Mn²⁺ were also added to provide a good living environment for microorganisms. After addition, the concentrations were: CuSO₄·5H₂O 9.68 μg/L, MgSO₄ 15.2 μg/L, MnCl₂·4H₂O 211 μg/L, FeCl₃ 5.88 μg/L, ZnSO₄·7H₂O 9.49 μg/L, CoCl₂·6H₂O 3.85 μg/L.
1.2.3 Experimental Inoculated Sludge
The inoculated sediment was taken from the Mihe River mud in Weifang City, Shandong Province, and the inoculated activated sludge was taken from the aeration tank of a wastewater treatment plant in Weifang City. The inoculated sediment was washed with distilled water and filtered through a 0.45 mm stainless steel sieve. The sediment and activated sludge were respectively laid at the bottom of the biofilm reactor at a ratio of 1:10.
1.3 Analysis Methods
1.3.1 Water Quality Indicator Analysis Methods
All water quality indicator analysis methods in the experiment followed national standard methods. The final value of each indicator was the average of three experiments. The specific measurement procedures are as follows:
pH, DO, T: Measured using a Mettler-Toledo dissolved oxygen meter. The specific steps were: ensure the meter is in good working condition; measure pH; measure dissolved oxygen (DO); measure temperature (T).
NH₄⁺-N: Measured by Nessler's reagent colorimetric method. The steps were: add an appropriate amount of sodium hydroxide solution to the water sample to convert NH₄⁺ to NH₃. Typically, 0.2 g NaOH is added per 10 mL of water sample. Add Nessler's reagent dropwise to the treated water sample to react with the generated NH₃ to form a yellow precipitate. Wait 15–30 minutes for full precipitate formation. The higher the NH₄⁺-N content, the darker the solution color. Calculate the NH₄⁺-N concentration based on the colorimetric results or absorbance readings.
COD: Measured by rapid digestion method. The steps were: add an appropriate amount of the water sample to a digestion flask, and add an appropriate amount of perchloric acid and sulfuric acid to achieve acidic digestion of the sample. Ensure the sample pH is below 2. Heat digestion: place the digestion flask in a constant temperature water bath or shaker heater, heat to boiling and maintain for a period (usually 2 hours). Cool the digestion flask to room temperature. Add indicator: add a small amount of silver sulfide as indicator to absorb the remaining chloric acid. Titration: use Rhodamine B for titration. Add Rhodamine B dropwise to the sample until the color changes from pink to yellow-green. Calculate the COD concentration.
1.3.2 Biofilm Characteristic Analysis Method
The biofilm amount was determined by the filling method. Polysaccharide content and protein content were determined by the nitric acid-distillation method and the Coomassie Brilliant Blue method using bovine serum albumin as the standard, respectively. The sum of the two was the EPS content. EPS content was used as an indicator for biofilm characteristic analysis. The experiment mainly analyzed the role of EPS in microbial cell aggregation. Biofilm samples before and after EPS removal were weighed, and 0.9% NaCl solution was added to prepare a suspension. The OD₆₀₀ value of the suspension was adjusted to 0.6 on a spectrophotometer. An appropriate amount of suspension was placed in a 5 mL centrifuge tube, and 3 mL of the supernatant was taken every 1 h to measure OD₆₀₀. The percentage of microbial cell aggregation at time t was calculated according to the formula.
The aggregation process followed a first-order kinetic equation.
2 RESULTS AND DISCUSSION
2.1 Analysis of Sewage Quality During Biofilm Attachment Start-up with Different Carriers
According to the operating parameters, the reactors were started, and the NH₄⁺-N and COD concentrations in the sewage were tested according to the procedures to compare the NH₄⁺-N and COD removal performance with different inoculation sources and different carriers.
From Figure 2(a), it can be seen that according to the NH₄⁺-N removal efficiency, the biofilm attachment start-up process was divided into three stages: adaptation, rapid growth, and stabilization. During the start-up process, the biofilm attachment efficiency of suspended and elastic carriers inoculated with activated sludge was better than that of carriers inoculated with sediment. Among them, reactor (b) achieved successful biofilm attachment on day 12, while reactor (a) was slightly slower. Due to its structural advantages, the elastic carrier required a shorter time for biofilm attachment, while the suspended carrier, due to its material, was more prone to biofilm detachment, reducing attachment efficiency. This indicates that the abundant active microorganisms in activated sludge improved the start-up efficiency of biofilm attachment.
From Figure 2(b), it can be seen that the COD removal rates of reactors (a)~(d) first increased and then stabilized. Among them, reactor (b) reached 80% removal faster than (a), while both (c) and (d) stabilized above 80% on day 9. This is because the strong adsorption capacity and cross-linked network structure of the elastic carrier increased the contact area between microorganisms and sewage, improving organic matter interception efficiency. In addition, the adhesiveness and richness of activated sludge also enhanced COD removal.
2.2 Analysis of the Influence of Different Factors on Biofilm Performance
Based on the above analysis results, the biofilm formed by reactor (d) (inoculated with activated sludge and elastic carrier), which had the best pollutant removal effect, was selected as the biofilm for subsequent experiments to further analyze the influence of different factors on biofilm performance. The results are shown in Figure 3.
2.2.1 DO
The experimental water temperature was set at 26°C–28°C, pH 6.8–7.2, and water flow rate 7.8–8.1 L/min. The NH₄⁺-N and COD removal rates were compared as the DO concentration increased from 1.5 mg/L to 4.5 mg/L, as shown in Figure 3(a).
From Figure 3(a), it can be seen that as DO concentration increased, the NH₄⁺-N and COD removal rates increased. When DO was 1.5 mg/L, the removal rates of NH₄⁺-N and COD were about 34% and 48%, respectively. At this time, the water body was in an anoxic state, with insufficient oxygen transfer to the biofilm, making it difficult for nitrifying bacteria to grow, resulting in low removal rates. Then, as DO increased to 3.5 mg/L, both removal rates increased significantly, and thereafter the change gradually flattened. That is, after DO exceeded 3.5 mg/L, DO was no longer a key factor affecting NH₄⁺-N and COD removal.
2.2.2 pH
The experimental water temperature was set at 26°C–28°C, water flow rate 7.8–8.1 L/min, and HCl and NaHCO₃ were used to adjust the pH. The NH₄⁺-N and COD removal rates were compared as pH increased from 6 to 9. It can be seen that as pH increased, the removal rates of NH₄⁺-N and COD first increased and then decreased, both reaching a maximum at pH 7. pH significantly affects nitrifying bacteria. The optimal pH for nitrifying bacteria growth is 7.5–8.0, and their activity is strongest at pH 7.0–7.8. Beyond this range, their activity drops sharply, leading to a decrease in NH₄⁺-N and COD removal rates, consistent with the figure.
2.2.3 Water Flow Velocity
The experimental water temperature was set at 26°C–28°C, pH 6.8–7.2, DO 3.5–4.5 mg/L. The NH₄⁺-N and COD removal rates were compared as the water flow velocity increased from 6.5 L/min to 9.5 L/min. It can be seen that before the flow velocity reached 8.0 L/min, the removal rates increased with increasing flow velocity, reaching a maximum at 8.0 L/min, and then began to decrease. This is because an appropriate flow velocity promotes water movement, accelerating the transport of substances and oxygen into the biofilm, but an excessively high flow velocity impacts the biofilm, causing some biofilm loss, thereby affecting the treatment efficiency.
2.2.4 COD Load
The experimental water temperature was set at 26°C–28°C, DO 3.5 mg/L, pH 7, water flow velocity 8 L/min. The NH₄⁺-N and COD removal rates were compared as the COD load increased from 15 g/m²·d to 30 g/m²·d. It can be seen that when the COD load increased from 15 g/m²·d to 30 g/m²·d, the removal rates increased rapidly and then gradually stabilized, reaching a maximum at a COD load of 25 g/m²·d. At this point, the NH₄⁺-N removal rate was close to 90%, and the COD removal rate exceeded 90%. When the COD load exceeded 25 g/m²·d, further increasing the load did not significantly change the removal rates. The optimal COD load for the biofilm used in the experiment was 25 g/m²·d.
2.2.5 Influent C/N Ratio
The experimental water temperature was set at 26°C–28°C, DO 3.5 mg/L, pH 7, water flow velocity 8 L/min. The NH₄⁺-N and COD removal rates were compared at influent C/N ratios of 5.3, 10.2, and 14.6, with results shown in Figure 3(e).
From Figure 3(e), it can be seen that when the influent C/N ratio was 5.3, the biofilm achieved a good NH₄⁺-N removal rate of 90%, but the COD removal rate was low. This is because a low C/N ratio favors the growth of autotrophic nitrifying bacteria, making them compete more strongly for nutrients and DO than heterotrophic nitrifying bacteria, leading to an imbalance between heterotrophic and autotrophic nitrifying bacteria, resulting in poor COD removal. When the influent C/N ratio was 10.2, the biofilm achieved relatively high removal rates for both NH₄⁺-N and COD, because under this condition, heterotrophic and autotrophic nitrifying bacteria grew in balance, and the biofilm purification efficiency was high. When the influent C/N ratio was 14.6, the COD removal rate was high, exceeding 90%, but the NH₄⁺-N removal rate dropped sharply to only about 65%. Therefore, when selecting the influent C/N ratio, both NH₄⁺-N and COD removal efficiencies should be considered.
3 CONCLUSION
To improve the effect of water pollution treatment, a bioaugmentation treatment method for urban water pollution based on biofilm was proposed. A biofilm reactor was constructed to compare the biofilm attachment start-up performance with different inoculation sources and different carriers, as well as the effects of DO, pH, water flow velocity, COD load, and influent C/N ratio on the sewage treatment efficiency of the biofilm. The analysis showed that the elastic carrier biofilm inoculated with activated sludge had higher removal rates for NH₄⁺-N and COD, achieving better treatment performance. Furthermore, when DO concentration was maintained at 3.5 mg/L, pH controlled within the suitable range of 7.0 to 7.8, water flow velocity stabilized at 8.0 L/min, COD load at 25 g/m²·d, and influent C/N ratio adjusted to 10.2, the biofilm exhibited the optimal pollutant removal effect. In summary, the bioaugmentation treatment method for urban water pollution based on biofilm shows significant advantages and potential. Generating biofilm tailored to local conditions for urban water pollution treatment provides ideas for improving urban water pollution and maintaining urban ecological balance.